Emerging Technologies And Biological Systems For Biogas Upgrading b514i

. Bauer et al., 2013b Bauer F, Persson T, Hulteberg C, Tamm D. Biogas upgrading: technology overview, comparison and perspectives for the future. Biofuels Bioprod Bioref. 2013b;7:499–511. Billet, 1995 Billet R. Packed Towers in Processing and Environmental Technology New York: VCH Publishers Inc.; 1995. Calix et al., 2008 Calix TF, Ferrentino G, Balaban MO. Measurement of highpressure carbon dioxide solubility in orange juice, apple juice, and model liquid foods. J Food Sci. 2008;79:E439–E445. Carroll and Mather, 1989 Carroll JJ, Mather AE. The solubility of hydrogen sulphide in water from 0 to 90 C and pressures to 1 MPa. Geochim Cosmochim Acta. 1989;55:1163–1170. Carroll and Mather, 1997 Carroll JJ, Mather AE. A model for the solubility of light hydrocarbons in water and aqueous solutions of alkanolamines. Chem Eng Sci. 1997;5:545–552.

Carroll et al., 1991 Carroll JJ, Slupsky JD, Mather AE. The solubility of carbon dioxide in water at low pressure. J Phys Chem Ref Data. 1991;20:1201–1209. Cavenati et al., 2005 Cavenati S, Grande CA, Rodrigues AE. Upgrade of methane from landfill gas by pressure swing adsorption. Energy Fuels. 2005;19:2545–2555. Chandra et al., 2012 Chandra R, Vijay VK, Subbarao PMV. Vehicular quality biomethane production from biogas by using an automated water scrubbing system. Int Sch Res Not. 2012;2012 https://doi.org/10.5402/2012/904167. Clever and Young, 1987 Clever HL, Young CL. Solubility Data Series, Methane Oxford: Pergamon Press; 1987;27–28. Cozma et al., 2013a Cozma P, Wukovits W, Mamaliga Friedl I, Gavrilescu M. Analysis and modelling of the solubility of biogas components in water for physical absorption processes. Environ Eng Manage J. 2013a;12:147–162. Cozma et al., 2013b Cozma P, Ghinea C, Mămăligă I, Wukovits W, Friedl A, Gavrilescu M. Environmental impact assessment of high-pressure water scrubbing biogas upgrading technology. Clean Soil Air Water. 2013b;41:917– 927. Cozma et al., 2014 Cozma P, Ghinea C, Mămăligă I, Wukovits W, Friedl A, Gavrilescu M. Modelling and simulation of high-pressure water scrubbing technology applied for biogas upgrading. Clean Technol Environ Policy. 2014;17:373–391. Duan et al., 2007 Duan Z, Sun R, Liu R, Zhu C. Accurate thermodynamic model for the calculation of H2S solubility in pure water and brines. Energy Fuels. 2007;21:2056–2065. Fernandes, 2011 Fernandes M. HETP evaluation of structured and randomic packing distillation column. Mass Transfer in Chemical Engineering Processes InTech 2011; Chapter 3, . Goffeng, 2013 Goffeng, B., 2013. Cryotechnology for biogas. . (Accessed June 2015).

Houghton, 2008 Houghton R. Emergency Characterization of Unknown Materials Taylor and Francis, CRC Press 2008. Huang et al., 1985 Huang SSS, Leu AD, Ng HJ, Robinson DB. The phasebehaviour of 2 mixtures of methane, carbon dioxide, hydrogen-sulfide, water. Fluid Phase Equilib. 1985;19:21–32. Hunter and Oyama, 2000 Hunter P, Oyama ST. Control of Volatile Organic Compound Emissions: Conventional and Emerging Technologies New York: John Wiley & Sons Inc.; 2000. IEA Bioenergy, 2013 IEA Bioenergy, 2013. Up-grading plant list. . (Accessed March 2015). Jonsson and Westman, 2011 Jonsson, S., Westman, J., 2011. Cryogenic Biogas Upgrading Using Plate Heat Exchangers (Master’s thesis within the Sustainable Energy Systems Master’s programme). Chalmers University of Technology, Goteborg, Sweden. Kapdi et al., 2005 Kapdi SS, Vijay VK, Rajesh SK, Prasad R. Biogas scrubbing, compression and storage: perspective and prospectus in Indian context. Renew Energy. 2005;30:1195–1202. Keskinen et al., 1990 Keskinen KI, Kinunen A, Nystrom L. Efficient approximate method for packed column separation performance simulation. Lapperanta Univ Technol. 1990;43:70–75. Kestin et al., 1978 Kestin J, Sokolov M, Wakeham WA. Viscosity of liquid water in the range −8°C to 150°C. J Phys Chem Ref Data. 1978;7. Kohl and Nielsen, 1997 Kohl A, Nielsen R. Gas Purification fifth ed. Houston, TX: Gulf Publishing Company; 1997. Läntelä et al., 2012 Läntelä J, Rasi S, Lehtinen J, Rintala J. Landfill gas upgrading with pilot-scale water scrubber: Performance assessment with absorption water recycling. Appl Energy. 2012;92:307–314. Lide, 1993 Lide DR. Handbook of Chemistry and Physics Boca Raton, FL: CRC Press; 1993.

Lin and Shyu, 1999 Lin SH, Shyu CT. Performance characteristics and modelling of carbon dioxide absorption by amines in a packed column. Waste Manage. 1999;19:255–262. McCabe et al., 1993 McCabe WL, Smith JC, Harriott P. Unit Operations in Mass Transfer McGraw Hill, Inc 1993. Morrison and Billett, 1952 Morrison TJ, Billett F. The salting-out of nonelectrolytes Part II The effect of variation in non-electrolyte. J Chem Soc. 1952;3819–3822. Namiot, 1961 Namiot AY. In: Clever HL, Young CL, eds. Solubility Data Series, Methane. Oxford: Pergamon Press; 1961;27–28. Namiot, 1967 Namiot AY. Reasons for small solubility of nonpolar gases in water. J Struct Chem. 1967;8:363–366. Ndiritu et al., 2013 Ndiritu HM, Kibicho K, Gathitu BB. Influence of flow parameters on capture of carbon dioxide gas by a wet scrubber. J Power Technol. 2013;95:9–15. Nozic, 2006 Nozic, M., 2006. Removal of carbon dioxide from biogas. . (Accessed January 2015). Patterson et al., 2011 Patterson T, Esteves S, Dinsdale R, Guwy A. An Evaluation of the policy and techno-economic factors affecting the potential for biogas upgrading for transport fuel use in the UK. Energy Policy. 2011;39:1806– 1816. Perry and Green, 1997 Perry RH, Green D. Perry’s Chemical Engineers’ Handbook New York: McGraw-Hill; 1997. Persson et al., 2006 Persson, M., Jonsson, O., Welliger, A., 2006. Evaluation of upgrading techniques to vehicle fuel standards and grid injection, IEA Bioenergy —Task 37-Energy From Biogas and Landfill Gas. IEA Bioenergy. . Pertl et al., 2010 Pertl A, Mostbauer P, Obersteiner G. Climate balance of biogas upgrading systems. Waste Manage. 2010;30:92–99.

Rasi et al., 2008 Rasi S, Lantela J, Veijanen A, Rintala J. Landfill gas upgrading with counter current water wash. Waste Manage. 2008;28:1528–1534. Rettich et al., 1981 Rettich TR, Handa YP, Battino R, Wilhelm E. Solubility of gases in liquids 13 High-precision determination of Henry’s constants for methane and ethane in liquid water at 275 to 328 K. J Phys Chem. 1981;85:3230–3237. Richardson et al., 2002 Richardson JF, Harker JH, Backhurst JR. Coulson and Richardson’s Chemical Engineering fifth ed. Oxford: Pergamon Press; 2002. Ryckebosch et al., 2011 Ryckebosch E, Drouillon M, Veruaeren H. Techniques for transformation of biogas to biomethane. Biomass Bioenergy. 2011;35:1633– 1645. Sander, 2011 Sander R. Carbon dioxide Henry’s Law data. NIST Chemistry Webbook National Institute of Standards and Technology 2011. Sen et al., 2013 Sen D, Sarkar S, Bhattacharjee S, Bandopadhya S, Ghosh S, Bhattacharjee C. Simulation of the effect of various operating parameters for the effective separation of carbon dioxide into an aqueous caustic soda solution in a packed bed using lattice Boltzmann simulation. Ind Eng Chem Res. 2013;52:1731–1742. Shao et al., 2012 Shao P, Dal-Cin M, Kumar A, Li H, Singh DP. Design and economics of a hybrid membrane–temperature swing adsorption process for upgrading biogas. J Membr Sci. 2012;413–414:17–18. Sherwood et al., 1938 Sherwood TK, Shipley GH, Holloway FAL. Flooding velocities in packed columns. Ind Eng Chem. 1938;30:765–769. Sloan and Koh, 2008 Sloan DE, Koh C. Clathrate Hydrates of Natural Gases third ed. Taylor and Francis, CRC Press 2008;120–121. Takahashi and Fujita, 1967 Takahashi T, Fujita KA. Correlation of flooding velocities in counter current gas-liquid or of column type. Mem Sch Eng Okayama Univ. 1967;2:43–49. Tan et al., 2012 Tan LS, Shariff AM, Lau KK, Bustam MA. Factors affecting CO2 absorption efficiency in packed column: a review. J Ind Eng Chem.

2012;18:1874–1883. Tippayawong and Thanompongchart, 2010 Tippayawong N, Thanompongchart P. Biogas quality upgrade by simultaneous removal of CO2 in a packed column reactor. Energy. 2010;35:4531–4535. Treybol, 1968 Treybol RE. Mass transfer operations third ed. New York: McGraw-Hill; 1968. Tynell et al., 2007 Tynell A, Borjesson G, Persson M. Microbial growth on pallrings – a problem when upgrading biogas with the technique absorption with water wash. Appl Biochem Biotechnol. 2007;141:299–319. Uttamote, 2011 Uttamote, T., 2011. Removal of CO2 and H2S From Biogas for Later Application in Solid Oxide Fuel Cell (Thesis. Master of Engineering, Sirindhorn International Institute of Technology), Thammasat University. . Vijay, 2007 Vijay, V.K., 2007. Biogas refining for production of bio-methane and its bottling for automotive applications and holistic development. In: Proceedings of International Symposium on EcoTopia Science. . Weiss, 1974 Weiss RF. Carbon dioxide in water and seawater: the solubility of a non-ideal gas. Mar Chem. 1974;2:203–215. Wellinger et al., 2013 Wellinger A, Murphy J, Baxter D. In: Hughes S, ed. The Biogas Handbook: Science, Production & Application Woodhead Publishing Series in Energy. Philadelphia, PA: Woodhead Publishing; 2013;2013. Wen and Hung, 1970 Wen WY, Hung JH. Thermodynamics of hydrocarbon gases in aqueous tetraalkyl ammonium salt solutions. J Phys Chem. 1970;74:170–180. Wilhelm et al., 1977 Wilhelm E, Battlno R, Wilcock RJ. Low-pressure solubility of gases in liquid water. Chem Rev. 1977;77. Xiao et al., 2014 Xiao Y, Yuan H, Pang Y, et al. CO2 removal from biogas by water washing system. Chin J Chem Eng. 2014;22:950–953.

Chapter 5

Recent developments in pressure swing adsorption for biomethane production

Goldy Shah¹, Shivali Sahota¹, Virendra Kumar Vijay¹, Kamal K. Pant² and Pooja Ghosh¹, ¹1Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India, ²2Department of Chemical Engineering, Indian Institute of Technology Delhi, New Delhi, India

Abstract

Different gas separation technologies such as absorption (both physical and chemical), pressure swing adsorption (PSA), membrane separation, and cryogenic treatment can be used to purify the CH4-rich biogas stream. PSA is the second-most commonly employed technique for biogas upgrading, with several commercial small-to-medium-scale plants existing worldwide. During the PSA process, biogas is compressed to a pressure between 4–10 bar and is fed to a vessel where it comes in to with a material (adsorbent) that selectively retains CO2. The adsorbent is a porous solid, usually with a high surface area. Recent research and experimental demonstrations have shown that adsorption technology paves the way for scaling up the biogas upgrading process with high energy efficiencies of up to 90%–95%. Moreover, PSA systems can be installed in any part of the world since they do not depend on temperature. Due to its low energy requirements, low capital costs, flexibility, and high efficiency in comparison to other gas separation methods, it has become one of the most preferred and widely implemented technologies.

Keywords

Pressure swing adsorption; biogas upgrading; biomethane; adsorbents

Chapter outline

Outline

5.1 Introduction 93

5.2 Types of swing adsorption technologies 95

5.2.1 Temperature swing adsorption 95 5.2.2 Electric swing adsorption 97 5.2.3 Vacuum swing adsorption 97 5.2.4 Pressure swing adsorption 98

5.3 Parameters influencing pressure swing adsorption 99

5.3.1 Process performance indicators 100 5.3.2 Design parameters 101

5.3.3 Adsorbents 103

5.4 Adsorption isotherm 107

5.5 Adsorption kinetics 109

5.5.1 Molecular diffusion 110 5.5.2 Knudsen diffusivity 110 5.5.3 Poiseuille diffusion or viscous diffusion 110

5.6 Mathematical modeling 112

5.7 Conclusion and future perspectives 113

References 113

5.1 Introduction

Discharge of carbon dioxide (CO2) into the environment s for 60% of the global warming caused by greenhouse gases. Over a 20-year period, methane is 84 times more potent than CO2 in of global warming potential, so there is a need to address both types of emissions for effectively reducing the impact of climate change (Weiland, 2010). Pressure swing adsorption (PSA) is a versatile and futuristic technology used for the cleaning and separation of a mixture of gases according to the adsorbent affinity and molecular characteristics. The PSA process is widely used for the purification and separation of a large number of industrial gas mixtures (Riboldi and Bolland, 2017). Starting from the patents of Guerin de Montgareuil (French Patent No. 1,233,261) and Skarstrom (US Patent No. 2,944,627) in 1957 and 1960, respectively, PSA has been applied to a variety of applications ranging from air drying and separation, hydrogen recovery and purification, CO2 capture, noble gas separation, hydrocarbon separation, n/iso paraffin separation, CH4 upgrading, and so forth. The chronological development of the PSA process is summarized in Fig. 5.1 (Ruthven et al., 1994). Recent literature on biogas upgradation suggests swing-based biogas purification technologies are progressively increasing for different applications.

Figure 5.1 Chronological order of the historical developments of the pressure swing adsorption (PSA) process.

There are several advantages to PSA, like low power consumption, low capital investment, flexibility to design, and high purification and regeneration efficiency, which make it a most favored and extensively used technology in industry for air separation, recovery of hydrogen, air drying, and lately for biogas upgradation. PSA processes are based not only on the capability of various adsorbent materials to selectively retain one or more components of a gas mixture under varying pressure conditions, but also on the basis of physical forces and molecular sizes of the gaseous components which decide their penetration and adsorption ability onto the adsorbent material (Jain et al., 2003). PSA is viewed as a promising technology due to its various advantageous features like cost-effectiveness, high regeneration, high productivity, chemicalfree usage, and low energy consumption. However, like all separation and purification processes, the adsorbent shows a very significant role and decides the efficiency of the process (Canevesi et al., 2019). Therefore, the major advantage of PSA in comparison to other adsorption processes, for example, temperature swing processes, is that pressure can be altered more quickly than temperature, thus making it possible to operate a PSA process at a much faster rate, thereby increasing the throughout per unit of adsorbent bed volume. However, the major limitation of the PSA process is that it is not feasible for less adsorbed species. However, if the favorably adsorbed component is too heavily adsorbed, then regeneration is done by a high vacuum pump for effective desorption.

5.2 Types of swing adsorption technologies

The term “swing” refers to a method with an adsorption and desorption cycle practiced by fluctuating operating parameters such as pressure, temperature, vacuum, and electric supply. It works on the adsorption principle in which a molecule of gas is selectively adsorbed on the surface of a solid. In general, adsorption is a surface phenomenon which involves selective attachment of one or more components of a mixture on the surface of a microporous solid, preferably with a large surface area per unit mass. The loaded particles are regenerated via a cyclic operation, for example, by decreasing the pressure to atmospheric pressure in the case of PSA, by applying a vacuum for desorbing material in the case of vacuum swing adsorption (VSA), by applying electric voltage in the case of electric swing adsorption (ESA), or by increasing the temperature at a constant level in temperature swing adsorption (TSA). The different types of swing adsorption technologies are discussed in detail below and Table 5.1 depicts the efficiency of different adsorption technologies.

Table 5.1

S. No.

Technology

Adsorbent

Methane purity (%)

Cycle time (s)

1

VPSA

Carbon molecular sieves (CMS-3K)

>98%

400

2

PSA

Zeolite 13X, CMS-3K, and MOF-508b

>98%

800

3

VPSA

Novel amine-based solid sorbent

>98%

900

4

PSA

Carbon molecular sieve

≤96%

840

5

PSA

Zeolite 5A

≥99%

500

6

PSA

Zeolite 13X

≥99%

800

7

VPSA

Zeolite 13X (CECA)

90%

615

8

PSA

Carbon molecular sieve (CMS)

≥97.8%

990

9

VPSA

Silica gel

98%

320

PSA, Pressure swing adsorption; VPSA, Vacuum pressure swing adsorption.

5.2.1 Temperature swing adsorption

When the regeneration of adsorbent is done by increasing the temperature of the adsorbent bed to remove the adsorbed gases this is described as TSA (Ntiamoah et al., 2016). Although the nature of a typical adsorption process is cyclic, commercial viability can be improved by out-of-phase product collection from several adsorbent beds, making the overall process pseudocontinuous. The TSA process comprises of the following three steps:

• Adsorption: At low temperature; • Desorption: By heating the bed to remove impurities; • Cooling: To return to the adsorption step.

TSA is more appropriate than PSA only when high methane purity is not attainable by PSA. Since pressure variations are not required in the TSA process, this decreases the power requirement and system complexity compared to PSA (Sonnleitner et al., 2018). However, the major disadvantage of using TSA is that it is very expensive when the temperature of adsorbent regeneration is higher than 400–450°C due to the difficulty in heating the low thermal conductivity adsorbent material to desorb impurities and regenerate the adsorbent. However, very limited work has been reported in the literature on the design of biogas purification systems by temperature swing adsorption (Ben-Mansour and Qasem, 2018).

5.2.2 Electric swing adsorption

When the regeneration of adsorbent is done by raising the temperature of the bed using a direct Joule effect low-voltage current to heat up the adsorbent, the process is termed ESA. In general, the principle of the ESA process operation is similar to TSA, but there are significant changes in their unit productivities to concentrate noncondensable gases. The major advantage of the ESA process is that it requires less heating time and has very high regeneration efficiency because it can rapidly heat up the adsorbent for regeneration (Petkovska et al., 2007). Hence, high purity, high throughput, and high recovery can be attained by ESA, as compared to TSA in which gases are used to heat the bed. The ESA process typically involves three steps, namely (1) feeding, (2) electrification, and (3) cooling for biogas upgradation. In the feed stage, the mixture of gases is fed to the adsorption tower at low temperature (Sullivan et al., 2004). Afterward, the heavy component is trapped in the adsorbent, while the component with less affinity is recovered at the top of the tower. The feed step stops just before the breakthrough. The feed step is followed by the electrification step, in which an electric current is supplied to the adsorbent bed, which causes an increase in the adsorbent temperature due to the Joule effect (Ribeiro et al., 2014). After removal of the adsorbate from the adsorption tower, cooling is employed to decrease the adsorbent temperature to recondition the adsorbent for the start of a new ESA cycle (Zhao et al., 2018).

5.2.3 Vacuum swing adsorption

The principle of VSA differs from pressure swing technology because it operates at near atmospheric temperature and pressure. In PSA, pressurized gas is used to feed and vent to an atmospheric pressure, but in the case of VSA, a vacuum blower is used to draw the gas (Shen et al., 2019). The combination of pressure

and vacuum is also used, that is, vacuum pressure swing adsorption (VPSA). In this method, pressurized gas is used to feed and a vacuum is used instead of purging. The advantage of using a vacuum step instead of a purge step is that by using the same pressure and product, the loss of methane during the purge step (where the less adsorbed species is used for purging) is always greater than the vacuum step (there is no need to use methane in a vacuum). The only option to save energy is possible if the step cycle is performed at high pressure, that is, above atmospheric pressure, and desorption is at a very low pressure (Cavenati et al., 2006).

5.2.4 Pressure swing adsorption

PSA is usually used to separate CO2 from biogas using physical properties. PSA is based on the principle of adsorption and is accomplished by an adsorbent on a porous bed material. PSA processes can be categorized according to the nature of the adsorption selectivity and whether the less strongly adsorbed species is recovered at high purity. The adsorption separation process mainly comprises of two steps: adsorption and desorption. In adsorption, the favorably adsorbed species is adsorbed by the adsorbent. In desorption, or regeneration, species are removed from the adsorbent so that adsorbent is ready to use again in the next cycle. The effluent during the adsorption step is the purified “raffinate” product from which the preferentially adsorbed species have been removed. The desorbate that is recovered during the regeneration step contains the more strongly adsorbed species in concentrated form and is called the “extract” product (Santos et al., 2011). The essential feature of PSA is that, in the regeneration step, adsorbent is regenerated by reducing the total pressure of the system, also to remove the adsorbed species (Wu et al., 2015). PSA processes operate at isothermal conditions, that is, at constant temperature, so the capacity is the difference in loading between adsorption in feed and the regeneration pressure. Before the more strongly adsorbed species break through the column, the feed step is usually stopped and, when the adsorption bed is fully desorbed, the regeneration step is terminated. Therefore, at a cyclic state, the process is always

at the mean position in the bed. The early PSA process was developed for equilibrium separations, primarily for the recovery of the less strongly adsorbed raffinate product at high purity (Heck et al., 2018). The early PSA process was based on two different techniques for regenerating the adsorbent and cleaning the void volume. The cycle developed by Skarstrom employed atmospheric desorption with product purge, where the air–liquid cycle utilized vacuum desorption. Evacuation to a very low absolute pressure may be necessary to achieve reasonable regeneration by vacuum desorption, especially when the isotherm for the more strongly adsorbed component is of a favorable form (Augelletti et al., 2017). However, vacuum desorption has other advantages and is still widely used, particularly for kinetic separations (Abdullah et al., 2019). The most common steps of the PSA cycle are as follows:

• Pressurization (with feed or raffinate product); • High-pressure feed with raffinate withdrawal; • Depressurization or blow down (cocurrent or countercurrent to the feed); • Desorption at the lower operating pressure; • Pressure equalization; • Rinse.

The various steps of the PSA process are described below in detail and are also shown in Fig. 5.2:

Figure 5.2 Schematic diagram showing the different steps of pressure swing adsorption (PSA).

Step 1: Feed/pressurization: In this step, the top end of the adsorption bed is kept closed, while biogas of a given concentration is fed to the bed for a certain period of time or until a certain desired pressure is reached. Step 2: Adsorption/elution: In this step the outlet of the bed is opened so as to collect the enriched biogas or biomethane (approx. 92%–99%). Step 3: Blowdown step: In this step, the inlet to the bed is kept closed, while the outlet is opened. As the operating pressure is higher than atmospheric pressure, the bed will experience a pressure drop at the outlet and hence the higher adsorbed component of the feed mixture will be released and the bed will be regenerated. The outlet opening can be done in many different ways, such as using different control valves. Step 4: Purge step: In this step, the bed is purged with feed biogas. Biogas is fed to the bed at the inlet while the outlet is left open to atmospheric pressure. This step is included to regenerate the bed further.

5.3 Parameters influencing pressure swing adsorption

The various processes and design parameters influencing PSA are discussed in detail below.

5.3.1 Process performance indicators

Undoubtedly, the PSA technology provides considerable environmental benefits, since it is based on the use of solid adsorbents which can be regenerated and does not involve the use of chemicals. Also, no heating is required for regeneration, thus energy requirements for the PSA process are low. The PSA units can be easily downscaled to skid-mounted modules suitable for small-scale applications. There are several general features of a PSA process that to a large extent explain both the advantages and limitations of the technology and hence determine the suitability for a given application (Shen et al., 2018):

where, NCH4 and NCO2 are the number of moles of CH4 and CO2 respectively, that come out of the column in the elution step.

where, NCH4 is as described above and FCH4 is the total number of moles of CH4 fed to the system in all of the steps of one cycle. Apart from these two there are two other major parameters that are used to evaluate the performance of a PSA system, namely productivity and power consumption.

where, FCH4 and Recovery are as described previously, and tcycle is the total cycle time in hours and Wtads is the total weight of the adsorbent in the bed. Power consumption is the amount of energy used by the system in order to produce the pressure swing effect or, in other words, the power used by the compressor to pressurize the system to the operating pressure. This can be calculated assuming adiabatic compression and using the following formula:

where is heat capacity ratio, R is universal gas constant, Tfeed is temperature of the feed, Phigh and Plow are pressures within which the system swings, Q is molar flow rate, and is mechanical efficiency. The major parameters and operating conditions that can be varied in order to maximize the aforementioned system purity and recovery are pressurization step time, adsorption step time, blowdown step time, purge step time, length of bed, diameter of bed, and bed porosity. The results of the manual optimization performed are given in the next section.

5.3.2 Design parameters

5.3.2.1 Pressure range

The term VSA is often used to denote a PSA cycle with desorption at subatmospheric pressure. This is a semantic choice. The performance of any PSA process is governed by the ratio of absolute (rather than gauge) pressures. The fact that desorption at subatmospheric pressure often leads to improved performance is due to the form of the equilibrium isotherm rather than to any intrinsic effect of a vacuum.

5.3.2.2 Pressure equalization

The invention of pressure equalization was the first improvement over Skarstrom’s cycle. After completion of the purge step in the first adsorption tower and the high-pressure adsorption step in the second adsorption tower,

rather than discharging the second bed directly, the two adsorption towers are connected to their exit end in order to balance or equalize the pressure. The gas exiting from the second tower thus partially pressurizes the first tower. After equalizing the pressure in both beds, they are disconnected and the process is reversed, that is, the first tower is pressurized with feed gas while the second tower starts the process of desorption (Grande and Rodrigues, 2007a,b). The advantage of using a pressure equalization step is to conserve energy, because the compressed gas from the high-pressure bed is used to partially pressurize the low-pressure bed. Since this gas is partially depleted of the strongly adsorbed species, the degree of separation is conserved and the blow down losses are reduced. It can be observed that the addition of a pressure equalization step favors product recovery but only in the case of a high-pressure swing. If the pressure swing is sufficiently low the addition of a pressure equalization step may be unworkable. Otherwise the pressure equalization step should always be integrated in the PSA process.

5.3.2.3 Time cycle

Pressurization time

It can be seen that as pressurization time increases there is a negligible effect on both purity and recovery until the breakthrough time is reached, beyond which there is a sharp decrease in both. This might mean that at the breakthrough time the bed might be getting completely saturated and beyond that it also becomes overcharged. Hence no further adsorption occurs and the additional feed is wasted, reducing both purity and recovery.

Adsorption time

As the elution time increases there is a negligible change in purity while recovery increases. This might mean that the breakthrough has not occurred yet, implying that all the adsorbent in the bed remains to be saturated. Hence whatever feed is being ed is purified to the highest purity possible.

Blowdown time

As blowdown time increases it can be seen that purity gradually increases while recovery also increases, but very slowly. As blowdown time increases, more opportunity exists for the higher adsorbed component (CO2) to escape from the adsorbent pellets, leading to greater regeneration of the bed (Canevesi et al., 2019). Therefore as the bed is more regenerated the purity is accordingly increased but recovery remains almost the same because neither the total amount of feed nor the moles of methane output change.

Purge time

As purge time increases purity increases and recovery decreases. As feed is being used to purge the bed a longer purge time implies more feed is being used, thereby purity increases but recovery decreases as this feed gas cannot be used as feed anymore.

5.3.2.4 Pressure swing adsorption sizing

The difference in adsorption capacity between the adsorption and the regeneration step is the main design criterion of PSA sizing. The column diameter is the first step in the design. The range of diameters usually selected is

between 3.5–4 m to ease the transport and handling of the large vessels (Santos et al., 2011). In this study, the typical superficial velocity varies from 0.01 to 0.05 m/s for the particular feed flow rate. The main factors which are responsible for the given number of moles of CO2 capture are the adsorption step time and the molar flow rate of the feed.

5.3.2.5 Pressure

The major goal of a PSA process is to control the pressure in the adsorption bed to obtain high purity and recovery. The basis of the selection of pressure depends on the equilibrium relationship. In an adsorption process, an isotherm shows the equilibrium loading of the species, which directly depends on the partial pressure of the species. The amount of fluid adsorbed is directly proportional to the increment of adsorption pressure; at high pressure the maximum amount of adsorbent is adsorbed. To find the pressure levels for adsorption and desorption processes, one should that the purity of the product increases with the adsorbate capacity of the adsorbent. For the constant selectivity process, high pressure causes the high adsorption of highly adsorbed gas and then the product purity will be increased. However, at high pressure, the less adsorbed gas is also adsorbed but it is always less in comparison to the more adsorbed gas, as long as the more adsorbed gas exists in high concentrations. For this reason, high pressure always leads to high purity for such systems.

5.3.2.6 Purge-to-feed ratio

In the PSA process, the purge step is a desorption process in which adsorbents are regenerated by using low pressure. Generally, from the product vessel with the raffinate at a high pressure, a fraction of the product stream is withdrawn to purge the bed and expended at a low pressure. The volume, that is, the purge to feed ratio required in the purge step, affects the product quality as well as its

recovery. Studies have reported that purity increases when the purge to feed ratio increases but decreases in the case of recovery (Jain et al., 2003).

5.3.2.7 Flow rate

Due to the feed flow rate increase, the purity of the adsorbed species also increases because, at constant time, an increased feed flow rate increases the adsorption zone, resulting in increased purity. Nevertheless, the recovery of CO2 decreases due to the large losses of the feed fraction.

5.3.2.8 Column length

Column length is an important factor when deg a PSA column. With an increase in column length, both the CO2 purity and recovery decrease. This is because with an increase in column length there occurs an increment in pressure drop leading to earlier breakthrough and degradation of the performance of the process.

5.3.3 Adsorbents

Different types of adsorbents can be used in PSA, namely zeolites, carbon-based adsorbent, metal organic frameworks (MOFs), and porous crystal. One possible reason for the limited range of adsorbents available is their pore diameter (3.2– 3.7 A), which is the most important criterion for the PSA technology for biogas upgrading (Uebbing et al., 2019). Nevertheless, any other adsorbent material can technically qualify for PSA application if the pore diameter is less than the

methane molecular diameter (3.8 Å) so that methane is not adsorbed inside the pores and at the same time it is greater than the CO2 molecule diameter (3.2 Å) so that the CO2 is adsorbed inside the pores. A promising adsorbent for application of swing adsorption processes should have properties such as a large working capacity, fast adsorption/desorption kinetics, high selectivity, and high stability (Yuan et al., 2013). The selection of adsorbent is a major and primary concern while deg any adsorption process. There are many adsorbents commercially available, but if the correct adsorbent is not chosen, the process will not deliver the required results. It is very important to develop a systematic technique to quickly evaluate the performance during the initial stages of a PSA system. Álvarez-Gutiérrez et al. (2018) developed a simplistic method that drives the decision-making process of adsorbent choice in a shorter time. The easiest way to assess adsorbents in any application is to run isotherm tests to obtain equilibrium of adsorption data (Álvarez-Gutiérrez et al., 2018). The most important physical properties to characterize an adsorbent are Brunauer-Emmett-Teller (BET) surface area and pore volume. The different types of adsorbents are described in detail below.

5.3.3.1 Carbon-based adsorbents

Carbon-based adsorbents are the most widely used adsorbents for CO2 capture because of their many advantages. For example, there is no need for pretreatment to remove moisture for separation, low polarity, low heat of adsorption which reduces the energy requirement for regeneration, and due to their low polarity behavior they adsorb more nonpolar compounds in comparison to other adsorbents (Siegelman et al., 2019). Carbon-based adsorbents are classified into four types as shown in Fig. 5.3. The most commonly used adsorbents among all carbon adsorbents are the carbon molecular sieves (CMSs) and activated carbon used for gas separation and biogas upgradation which are discussed below.

Figure 5.3 Carbon-based adsorbents.

Activated carbons

Activated carbon is produced in many different forms that differ mainly in pore size distribution and surface polarity. The nature of the final product depends on both the starting material and the activation procedure. For liquid-phase adsorption a relatively large pore size is required, and such materials can be made by both thermal and chemical activation procedures from a wide range of carbonaceous materials. The activated carbons used in gas adsorption generally have much smaller pores, with a substantial fraction of the total porosity in the micropore range. These adsorbents are generally made by thermal activation from a relatively dense form of carbon such as bituminous coal (Sarker et al., 2017). High-area small-pore carbons may also be made from sources such as coconut shells, but the product generally has insufficient physical strength for PSA applications. Regeneration can be laborious and energy intensive because it is done at high temperatures. However, it is more economical to purchase a new activated carbon material from a supplier than from onsite regeneration (Tagliabue et al., 2009). Activated carbons are used in the natural gas industry for desulfurization and dehydration.

Carbon molecular sieves

CMSs are among the most commonly used adsorbents for biogas upgrading. CMSs are excellent materials to separate different gaseous components present in biogas. The gaseous component molecules are loosely adsorbed in the pores of the carbon sieve but not irreversibly bound. The selectivity of adsorption is achieved by different mesh sizes and/or use of different gas pressures. Carbon dioxide separation from natural gas by CMS-based PSA systems has been widely studied, and commercial units (suitable also for nitrogen rejection) are

available. The preparation of CMS is comparable to activated carbons but includes additional treatment with organic species that are cracked or polymerized on the carbonized matter. In many liquid-phase applications, activated carbon is used in powder form, but for gas-phase applications, larger particles are needed (Sarker et al., 2017). These are made either directly by crushing and screening or more commonly by granulation of the powder using binders such as pitch, which can be activated to some extent during the final thermal treatment. The preparation of activated carbon in fiber form is a relatively new development which holds considerable promise for the future. The diameter of the fibers is small leading to reduction in diffusional resistance. The preparation of CMSs is broadly similar to activated carbon but often includes an additional treatment with species such as benzene or acetylene that are easily polymerized or cracked on the surface. By careful control of the condition a very uniform pore size is achieved. The pore size distribution of CMS is maintained within a very narrow range by specific control of synthesis conditions. The basis of the separation mechanism from activated carbon is the very fine pore size tuning (Reid and Thomas, 1999). CMSs are generally used as a substitute for zeolite and activated carbon. Zeolite and activated carbon are equilibrium-based adsorbents which are based on the capacity to adsorb carbon dioxide, whereas CMSs are kinetic-based adsorbents that have micropores allowing contaminant molecules to penetrate faster than methane.

5.3.3.2 Zeolites

In contrast to the other adsorbents, zeolites are crystalline in nature with uniform pore sizes and dimensions. Zeolites are polar and the micropores are actually intracrystalline channels with dimensions precisely determined by the crystal structure. Therefore there is no virtual distribution of micropore size, and these adsorbents show well-defined size-selective molecular sieve properties, exclusion of molecules larger than a certain critical size, and strong restriction of diffusion of molecules. The most commonly used zeolites are zeolite A, zeolite X or Y, and ZSM-5. In zeolite A, there are three types of sites (synthetic,

hydrous, and alkali aluminosilicates), so in the Ca²+ form, all cations can be accommodated in the type I sites, that is, they are synthetic where they do not obstruct the channels (Zabielska et al., 2018). The effective dimension of the channel is then limited by the aperture of the eight-membered oxygen rings and has a free diameter of about 4.3 Å. Since molecules with diameters up to about 5 Å can penetrate, this is referred to as a 5 Å sieve. The framework structures of X and Y zeolites are the same and these materials differ only in the Si-to-Al ratio and therefore in the number of exchangeable cations (Mosca, 2009). The pore structure is very open, with construction being twelve-membered oxygen rings with a free diameter of up to 7.5 Å. Molecules with diameter up to about 8.5 Å can penetrate these channels with little hindrance, and this includes all common gaseous species. ZSM-5 and silicate are essentially the same material with a high silica structure. ZSM-5 normally contains measurable aluminum with a corresponding proportion of cations. The pore network is three-dimensional, and the dimensions of the channels are limited by ten-membered oxygen rings with a free aperture of about 6 Å (Shokroo et al., 2016).

5.3.3.3 Porous crystals

MOFs are crystalline compounds with two- or three-dimensional porous structures consisting of metal ions or clusters coordinated to organic molecules. In some cases, the pores can be used for the storage of gases such as CO2. MOFs can be synthesized with many of the functional competences of zeolites (Tagliabue et al., 2009). They are formed due to the coordination bonds between metal salt and multidentate ligands. MOF solids are highly robust materials provided by the strong bonding of inorganic molecules and linking organic units (Pirngruber et al., 2012). MOFs are promising candidates as separation materials for CO2 capture. Several MOFs have been proposed as adsorbents for CO2 separation processes, and among these Cu-BTC [polymeric copper (II) benzene1,3,5-tricarboxylate] proved to be endowed with CO2 adsorption performances that are higher than those of typical adsorbents such as 13X zeolite. The distribution of pore size is classified into three major classes by size:

• Micropores<20 Å; • Mesopores 20–500 Å; • Macropores>500 Å.

In a micropore, the guest molecule never escapes from the field of the solid surface without regeneration, even at the center of the pore. It is therefore reasonable to consider all molecules within a micropore to be in the “adsorbed” phase. By contrast, in mesopores and macropores, the molecules in the central region of the pore are essentially free from the force field of the surface; therefore it becomes physically reasonable to consider the pore as a two-phase system containing both adsorbed molecules at the surface and free gaseous molecules in the central region. Thus, the distinction between a micropore and mesopore really depends on the ratio of pore diameter to molecular diameter rather than on absolute pore size. Nevertheless, for PSA processes that deal in general with relatively small molecules, 20 Å is a reasonably common choice. Macropores contain much less surface area in comparison to the pore volume and so contribute little to the adsorptive capacity. Their main function is to facilitate transport (diffusion) within the particle by providing a network of superhighways to allow molecules to penetrate rapidly into the interior of the adsorbent particle.

5.4 Adsorption isotherm

Adsorption is generally studied/modeled by concentration curves known as adsorption isotherms. If the adsorbate and adsorbent are left in for enough time, the adsorbate adsorbed on the adsorbent surface and adsorbate in the solution (gas mixture) will reach equilibrium and this relationship is what is studied by adsorption isotherms. An adsorption isotherm is the equation (curve) relating the equilibrium concentration of the adsorbate on the adsorbent surface to the concentration of the adsorbate in the fluid that is in with the adsorbent. Adsorption isotherms also give the amount of adsorbate adsorbed onto the surface and the equilibrium concentration of the adsorbate in the fluid at a given temperature. Many different mathematical forms of isotherms have been developed over the years (Chahbani and Tondeur, 2000) as shown in Table 5.2 and described in brief below:

1. Henry’s (linear) isotherm: It is the most common and widely used form of isotherm. In this form the assumption is that the amount of adsorbate adsorbed on the surface is proportional to the partial pressure of the adsorbent or the concentration of adsorbent in the fluid in with the surface.

where q is the amount of gas “i” adsorbed per unit gram of adsorbent, k is Henry constant, and c is the concentration of gas. 2. Langmuir isotherm: This assumes that the adsorbate gas behaves as an ideal gas under isothermal conditions (Ho, 2006). Further more, the Langmuir model is derived by considering the following assumptions

a) Monolayer coverage; b) Equilibrium model; c) All adsorption sites are equally probable; d) Adsorbate binding is treated as a second-order reaction between the empty surface site and the adsorbate molecule .

The equation derived from the previous assumptions takes the form:

where q is the amount of gas “i” adsorbed per unit gram of adsorbent, q* is the equilibrium adsorbed phase concentration, and KLC is the Langmuir constant. 3. Freundlich isotherm: This isotherm is a completely empirical model. It is best used when the isothermal data cannot be fit by other theoretically based isotherm models. The equation is of the form:

where q is the amount of gas “i” adsorbed per unit gram of adsorbent, KF is Freundlich constant, and Cn is the concentration of gas. 4. BET isotherm: This isotherm describes multilayer adsorption. The assumptions considered are:

a) Gas molecules adsorb onto the surface of the adsorbent in multiple layers infinitely; b) There is no interaction between each adsorption layer; c) Langmuir isotherm applies to each layer and no transmigration occurs between the layers; d) There is equal energy of adsorption for each layer except the first layer.

Table 5.2

Isotherm Langmuir Freundlich Temkin Brunauer-Emmett-Teller (BET)

Equation

Plot

The BET isotherm is of the form:

where v is the adsorbed gas quantity, p0 is the saturation pressure of adsorbate, p is the equilibrium pressure of adsorbate, c is the BET constant=exp (E1 – E2 /RT), E1 is the heat of adsorption for the first layer, and E2 is the heat of vaporization. A favorable isotherm curve is defined as a curve which is convex in curvature and also, in the isotherm, the adsorbed phase is higher than the fluid-phase concentration (Mosca, 2009). However, once the isotherm is favorable for an adsorption stage it is unfavorable for a desorption stage, there after the start and end steps are reversed and the opposite applies.

5.5 Adsorption kinetics

Adsorption kinetics refers to the rate of adsorption and desorption, and there are mathematical models that describe this behavior. Adsorption kinetic models are usually categorized into adsorption reaction models and adsorption diffusion models (Haghpanah et al., 2013). The adsorption reaction models are based on chemical reaction kinetic models, and take the entire adsorption process as a whole and do not consider the different steps involved (Qiu et al., 2009). The adsorption diffusion models are dependent on three different steps occurring during adsorption:

• External or film diffusion: this is defined as the diffusion taking place across the fluid film surrounding the adsorbent particle; • Internal or intraparticle diffusion: this is defined as the diffusion taking place from the particle surface to the pellet interior; • Mass action: in this type of model, adsorption and desorption take place between the active sites and the adsorbate.

Similar to adsorption isotherms, different mathematical forms of kinetics rate equations have been developed. Generally adsorbent pellets are made by compressing micro- or mesoporous materials into pellets of different shapes. Hence there are not only micro/mesopores in the pellets but also macropores. The different phases involved in the diffusion of gas in an adsorbent bed are shown in Fig. 5.4 and discussed below.

Figure 5.4 Phases of diffusion of gas from bulk to solid.

In the micro- or mesopores, first, the gas occupies interstitial sites between the adsorbent pellets by axial dispersion. Then it forms a film around the adsorbent pellets and film diffusion transfer occurs from the bulk of the gas (from interstitial sites) into the pellets. Once inside the pellets, the gas molecules diffuse through the micropores in order to enter the adsorbent particles, where they can find mesopores (Ackley and Yang, 1990). Then, finally, once they diffuse through the micropores they find the adsorbent surface to which they are adsorbed. In the macropores diffusion occurs via one or a combination of three diffusion mechanisms, namely, molecular diffusion, Knudsen diffusion, or viscous (Poiseuille) diffusion. Each of these is discussed briefly below.

5.5.1 Molecular diffusion

When the size of the macropores is greater than the molecular size of the gas molecules, that is, the mean free path of the species being measured, diffusion of the gas happens through random collisions between the gas molecules. Molecular diffusivity of a binary mixture of gases can be estimated using the Chapman-Enskog equation:

where M1 and M2 are the molecular weights of gases 1 and 2, respectively, T is the temperature (K), and p is the total pressure (atm).

5.5.2 Knudsen diffusivity

When the mean free path of the molecular species being considered is greater than the pore diameter then the diffusion occurs through this mechanism, wherein the diffusion of the gas occurs through the collision of the molecules with the macropore walls. The pure Knudsen diffusion coefficient can be calculated using the following equation:

where dp is the pore diameter (m), R is the gas constant (J/kmol K), T is the temperature (K), MA is the molecular weight (kg/kmol), and DA is the diffusivity (m²/s).

5.5.3 Poiseuille diffusion or viscous diffusion

This kind of diffusion is encountered when the flow within the macropores is due to the net pressure gradient. Hence, the gas moves in the pores with with the walls and as a continuous medium, similar to flow in a capillary. This type of diffusivity is generally negligible in packed beds. The Poiseuille diffusivity coefficient can be estimated by using the following equation:

Further, when the mean free path of the diffusing gas species is comparable to the pore diameter, diffusion occurs through both the molecular and Knudsen mechanisms, thereby the effective diffusivity coefficient in such a case is given by the following equation:

All the diffusivity coefficients mentioned above are calculated considering smooth nonporous pore walls, which is never the situation, hence a tortuosity and porosity factor always needs to be considered, which is:

Diffusion through micropores has been modeled using many models. In these models three major transport mechanisms have been identified: (1) pore model, wherein the diffusion resistance is distributed in the micropore interior; (2) barrier model, wherein the diffusion resistance is at the pore mouth; and (3) a dual resistance model, equivalent to the Bi-LDF model, which is a combination of the first two. All use either the Fickian diffusion law or the Maxwell-Stefan diffusion law to solve for mass fluxes. There are other models such as Weber-Morris model, Dunwald-Wagner model, and homogeneous solid diffusion model that have been developed to rigorously solve for mass balances of intraparticle diffusion, that is, microporous diffusion (Medrano et al., 2019). Given the experimental loading data, these models can be regressed and fit to get the appropriate micropore diffusivity. Owing to the difficulty in studying, modeling, and solving, micropore diffusion models have been developed to include overall mass transfer resistances, including:

• Solid diffusion model; • Linear driving force (LDF) model; • Vermeulen model; • Nakao and Suzuki model.

The most popular among the above-mentioned models is the LDF model created by Glueckauf and Coates which is discussed here. The LDF model for the adsorption kinetics is often used for estimation of adsorption column dynamic data. The LDF equation approximates the diffusion-controlled kinetics for many adsorption systems up to a satisfactory level.

5.6 Mathematical modeling

Since the start of the use of the PSA process, a wide variety of mathematical models have been suggested, based on theories that extend from simple to complex. To quantify the system and study the variables, a mathematical model for the system is developed in which physical parameters and variables are used to express the physical phenomena occurring in the adsorbent bed in the form of mathematical equations (Nikolaidis et al., 2017). These mathematical models can be solved to study the variables and mechanistic changes in space and/or time. Nevertheless, all models are based on certain assumptions as real system modeling is very complex and tedious. Adsorption is generally favored at low temperature and decreases by increasing the temperature (Li et al., 2014). Therefore the PSA process is assumed to be under isothermal conditions. The PSA process is described by a set of partial differential equations and linear equations for understanding the mass, energy, and momentum balance of the process (Arya et al., 2015; Canevesi et al., 2019; Bokare et al., 2020). The model equations used are described below (Table 5.3).

Table 5.3

Symbol

Unit

Meaning

C

mol/cm³

Total concentration in gas phase

qi

mol/g-adsorbent

Amount of gas ‘i’ adsorbed per unit gram of adsorbent

ρs

gads/cm³

Adsorbent density

Ci

mol/cm³

Concentration in gas phase

Dax

cm²/s

Axial dispersion coefficient

qi*

mol/g-adsorbent

Equilibrium adsorbed phase concentration

Overall material balance:

Component material balance (two equations, one for CO2 and the other for CH4):

Mass transfer rate (LDF equation) (one for each component):

Equilibrium loading equation (Langmuir) (one for each component):

Ergun equation (for pressure drop in the bed):

Ideal gas equation:

5.7 Conclusion and future perspectives

PSA is a promising technology for biogas upgrading due to the low energy requirements, safety, flexibility of design, and high efficiency in comparison to other gas separation methods. Compared to other adsorption processes such as TSA, PSA offers the advantages of faster operation rate and higher throughout per unit of adsorbent bed volume. However, its major limitation is that its application is limited to those components only that are not too strongly adsorbed. This is because, if the preferentially adsorbed species is too strongly adsorbed, an uneconomically high vacuum is required for desorption during the regeneration step. Thus, for very strongly adsorbed components, thermal swing is generally the preferred option. Most of the R&D has focused on the design of promising novel adsorbents such as MOFs and zeolitic imidazolate frameworks. More R&D needs to be directed toward minimizing the number of PSA units, developing new performance indicators for the evaluation of adsorbent efficiency, and segregating H2S and CO2 in a single column.

References

Abdullah et al., 2019 Abdullah A, Idris I, Shamsudin IK, Othman MR. Methane enrichment from high carbon dioxide content natural gas by pressure swing adsorption. J Nat Gas Sci Eng. 2019;69:102929. Ackley and Yang, 1990 Ackley MW, Yang RT. Kinetic separation by pressure swing adsorption: method of characteristics model. AIChE J. 1990;36(8):1229– 1238. Álvarez-Gutiérrez et al., 2018 Álvarez-Gutiérrez N, Gil MV, Rubiera F, Pevida C. Simplistic approach for preliminary screening of potential carbon adsorbents for CO2 separation from biogas. J CO2 Utilization. 2018;28:207–215. Arya et al., 2015 Arya A, Divekar S, Rawat R, et al. Upgrading biogas at low pressure by vacuum swing adsorption. Ind & Eng Chem Res. 2015;54(1):404– 413. Augelletti et al., 2017 Augelletti R, Conti M, Annesini MC. Pressure swing adsorption for biogas upgrading A new process configuration for the separation of biomethane and carbon dioxide. J Clean Prod. 2017;140:1390–1398. Ben-Mansour and Qasem, 2018 Ben-Mansour R, Qasem NA. An efficient temperature swing adsorption (TSA) process for separating CO2 from CO2/N2 mixture using Mg-MOF-74. Energy Convers Manag. 2018;156:10–24. Bokare et al., 2020 Bokare M, Bano S, Antony PS, Pal S, Biniwale R. Selective separation of carbon dioxide from biogas mixture using mesoporous ceria and zirconium hydroxide. Adsorption. 2020;26(1):51–59. Canevesi et al., 2019 Canevesi RL, Andreassen KA, Silva EA, Borba CE, Grande CA. Evaluation of simplified pressure swing adsorption cycles for biomethane production. Adsorption. 2019;25(4):783–793. Cavenati et al., 2006 Cavenati S, Grande CA, Rodrigues AE. Removal of carbon

dioxide from natural gas by vacuum pressure swing adsorption. Energy & fuels. 2006;20(6):2648–2659. Chahbani and Tondeur, 2000 Chahbani MH, Tondeur D. Mass transfer kinetics in pressure swing adsorption. Sep Purif Technol. 2000;20(2-3):185–196. Fatehi et al., 1995 Fatehi AI, Loughlin KF, Hassan MM. Separation of methane —nitrogen mixtures by pressure swing adsorption using a carbon molecular sieve. Gas Sep Purif. 1995;9(3):199–204. Fujiki et al., 2017 Fujiki J, Chowdhury FA, Yamada H, Yogo K. Highly efficient post-combustion CO2 capture by low-temperature steam-aided vacuum swing adsorption using a novel polyamine-based solid sorbent. Chem Eng J. 2017;307:273–282. Grande and Rodrigues, 2007a Grande CA, Rodrigues AE. Biogas to fuel by vacuum pressure swing adsorption I Behavior of equilibrium and kinetic-based adsorbents. Ind & Eng Chem Res. 2007a;46(13):4595–4605. Grande and Rodrigues, 2007b Grande CA, Rodrigues AE. Layered vacuum pressure-swing adsorption for biogas upgrading. Ind Eng Chem Res. 2007b;46(23):7844–7848. Haghpanah et al., 2013 Haghpanah R, Majumder A, Nilam R, et al. Multiobjective optimization of a four-step adsorption process for postcombustion CO2 capture via finite volume simulation. Ind & Eng Chem Res. 2013;52(11):4249–4265. Heck et al., 2018 Heck HH, Hall ML, dos Santos R, Tomadakis MM. Pressure swing adsorption separation of H2S/CO2/CH4 gas mixtures with molecular sieves 4A, 5A, and 13X. Sep Sci Technol. 2018;53(10):1490–1497. Ho, 2006 Ho YS. Review of second-order models for adsorption systems. J Hazard Mater. 2006;136(3):681–689. Jain et al., 2003 Jain S, Moharir AS, Li P, Wozny G. Heuristic design of pressure swing adsorption: a preliminary study. Sep Purif Technol. 2003;33(1):25–43. Li et al., 2014 Li Z, Xiao G, Yang Q, Xiao Y, Zhong C. Computational exploration of metal–organic frameworks for CO2/CH4 separation via

temperature swing adsorption. Chem Eng Sci. 2014;120:59–66. Medrano et al., 2019 Medrano JA, Llosa-Tanco MA, Tanaka DAP, Gallucci F. Membranes utilization for biogas upgrading to synthetic natural gas. In Substitute Natural Gas from Waste Academic Press 2019;245–274. Mosca, 2009 Mosca, A., 2009. Structured zeolite adsorbents for PSA applications (Doctoral dissertation, Luleåtekniskauniversitet). Nikolaidis et al., 2017 Nikolaidis GN, Kikkinides ES, Georgiadis MC. Modelling and optimization of pressure swing adsorption (PSA) processes for post-combustion CO2 capture from flue gas. Process Syst Mater CO2 Capture: Modelling, Design, Control Integr. 2017;343. Ntiamoah et al., 2016 Ntiamoah A, Ling J, Xiao P, Webley PA, Zhai Y. CO2 capture by temperature swing adsorption: use of hot CO2-rich gas for regeneration. Ind & Eng Chem Res. 2016;55(3):703–713. Petkovska et al., 2007 Petkovska M, Antov-Bozalo D, Markovic A, Sullivan P. Multiphysics modeling of electric-swing adsorption system with in-vessel condensation. Adsorption. 2007;13(3-4):357–372. Pirngruber et al., 2012 Pirngruber GD, Hamon L, Bourrelly S, et al. A method for screening the potential of MOFs as CO2 adsorbents in pressure swing adsorption processes. ChemSusChem. 2012;5(4):762–776. Qiu et al., 2009 Qiu H, Lv L, Pan BC, Zhang QJ, Zhang WM, Zhang QX. Critical review in adsorption kinetic models. J Zhejiang University-Science A. 2009;10(5):716–724. Reid and Thomas, 1999 Reid CR, Thomas KM. Adsorption of gases on a carbon molecular sieve used for air separation: linear adsorptives as probes for kinetic selectivity. Langmuir. 1999;15(9):3206–3218. Ribeiro et al., 2014 Ribeiro RPPL, Grande CA, Rodrigues AE. Electric swing adsorption for gas separation and purification: a review. Sep Sci Technol. 2014;49(13):1985–2002. Riboldi and Bolland, 2017 Riboldi L, Bolland O. Overview on pressure swing adsorption (PSA) as CO2 capture technology: state-of-the-art, limits and

potentials. Energy Procedia. 2017;114:2390–2400. Ruthven et al., 1994 Ruthven, D.M. and Shamsuzzaman, F., Kent, S.K., 1994. Pressure swing adsorption. New York: VCH Publishers, 1(994), p. 235. Santos et al., 2011 Santos MP, Grande CA, Rodrigues AE. Pressure swing adsorption for biogas upgrading Effect of recycling streams in pressure swing adsorption design. Ind & Eng Chem Res. 2011;50(2):974–985. Sarker et al., 2017 Sarker AI, Aroonwilas A, Veawab A. Equilibrium and kinetic behaviour of CO2 adsorption onto zeolites, carbon molecular sieve and activated carbons. Energy Procedia. 2017;114:2450–2459. Shen et al., 2018 Shen Y, Shi W, Zhang D, Na P, Fu B. The removal and capture of CO2 from biogas by vacuum pressure swing process using silica gel. J CO2 Utilization. 2018;27:259–271. Shen et al., 2019 Shen, Y., Shi, W., Zhang, D., Na, P. and Tang, Z., 2019. Recovery of light hydrocarbons from natural gas by vacuum pressure swing adsorption Dube, O., Celik, C.E. and McNamara, T.A., Praxair Technology Inc, 2019. Control of swing adsorption process cycle time with ambient CO2 monitoring. U.S. Patent 10,427,090.process. Journal of Natural Gas Science and Engineering, 68, p. 102895. Shokroo et al., 2016 Shokroo EJ, Farsani DJ, Meymandi HK, Yadollahi N. Comparative study of zeolite 5A and zeolite 13X in air separation by pressure swing adsorption. Korean J Chem Eng. 2016;33(4):1391–1401. Siegelman et al., 2019 Siegelman RL, Milner PJ, Kim EJ, Weston SC, Long JR. Challenges and opportunities for adsorption-based CO2 capture from natural gas combined cycle emissions. Energy & Environ Sci. 2019;12(7):2161–2173. Sonnleitner et al., 2018 Sonnleitner E, Schöny G, Hofbauer H. Assessment of zeolite 13X and Lewatit® VP OC 1065 for application in a continuous temperature swing adsorption process for biogas upgrading. Biomass Convers Biorefinery. 2018;8(2):379–395. Sullivan et al., 2004 Sullivan PD, Rood MJ, Grevillot G, Wander JD, Hay KJ. Activated carbon fiber cloth electrothermal swing adsorption system. Environ Sci & Technol. 2004;38(18):4865–4877.

Tagliabue et al., 2009 Tagliabue M, Farrusseng D, Valencia S, et al. Natural gas treating by selective adsorption: material science and chemical engineering interplay. Chem Eng J. 2009;155(3):553–566. Uebbing et al., 2019 Uebbing J, Rihko-Struckmann LK, Sundmacher K. Exergetic assessment of CO2 methanation processes for the chemical storage of renewable energies. Appl Energy. 2019;233:271–282. Weiland, 2010 Weiland P. Biogas production: current state and perspectives. Appl microbiology Biotechnol. 2010;85(4):849–860. Wu et al., 2015 Wu B, Zhang X, Xu Y, Bao D, Zhang S. Assessment of the energy consumption of the biogas upgrading process with pressure swing adsorption using novel adsorbents. J Clean Prod. 2015;101:251–261. Yuan et al., 2013 Yuan B, Wu X, Chen Y, Huang J, Luo H, Deng S. Adsorption of CO2, CH4, and N2 on ordered mesoporous carbon: approach for greenhouse gases capture and biogas upgrading. Environ Sci & Technol. 2013;47(10):5474– 5480. Zabielska et al., 2018 Zabielska K, Aleksandrzak T, Gabruś E. Adsorption equilibrium of carbon dioxide on zeolite 1X at high pressures. Chem Process Eng. 2018;39(3):309–321. Zhao et al., 2018 Zhao R, Liu L, Zhao L, Deng S, Li H. Thermodynamic analysis on carbon dioxide capture by electric swing adsorption (ESA) technology. J CO2 Utilization. 2018;26:388–396.

Chapter 6

Membrane-based technology for methane separation from biogas

Birgir Norddahl, M.C. Roda-Serrat, M. Errico and K.V. Christensen, Department of Green Technology, Faculty of Engineering, University of Southern Denmark, Odense M, Denmark

Abstract

In this chapter a historical and technical background for membrane gas separation is given, together with a review of the basic commonly used in the study of membrane separation technology. The mechanism of mass transport through the membrane, how this depends on the chosen membrane material, and how this affects the separation of gas molecules in biogas upgrading is described in detail. The present industrial membrane material of choice for biogas upgrading is polyimide, while carbon molecular sieves and mixed-matrix membranes combining polymers with either zeolites, graphene oxide, or metal-organic frameworks are promising emerging membrane materials. Biogas upgrading requires multistage membrane systems, preferably combined with cryogenic distillation to obtain grid gas and highpurity CO2. The economic evaluation indicates that production of grid-quality gas and CO2 from biogas could be economically attractive.

Keywords

Biogas upgrading; biomethane; membrane technology; gas separation

Chapter outline

Outline

6.1 Introduction: how the basic membrane processes for gas separation have evolved 117

6.2 Basic of gas separation on membranes 120

6.3 Membrane materials and structures 127

6.3.1 Polymer structures and their influence in permeation 127 6.3.2 Inorganic membranes for gas separation 132 6.3.3 Carbon molecular sieve membranes 132 6.3.4 Mixed-matrix membranes 134 6.3.5 Results of membrane operations with different materials 135

6.4 Theory of transport in gas separation on membranes 135

6.4.1 Transport through rubbery polymers 135 6.4.2 Transport equations through glassy polymers 137

6.5 Membrane configurations and plant design for upgrading biogas 140

6.6 Recent developments in membrane-based CO2/CH4 separation 142

6.6.1 Biogas upgrading by cryogenic and hybrid cryogenic-membrane separation 143 6.6.2 Biogas upgrading by absorption and hybrid absorption-membrane 145 6.6.3 Microbial conversion of CO2 to CH4 on a membrane diff 148

6.7 Summary and outlook 150

6.8 Future developments 151

References 152

6.1 Introduction: how the basic membrane processes for gas separation have evolved

Historically, gas separation through membranes or diaphragms has been studied since Thomas Graham measured permeation rates on a number of gases under different conditions. Those experiments led to the Graham diffusion theory back in 1833 (Graham, 1833). However, only in the last 30 years has gas separation on membrane materials reached industrial relevance. The first major application of membranes for gas separation was probably the use of microporous metal membranes for the separation of U²³⁵ from U²³⁸ in the Manhattan project, which led to the production of the first nuclear bomb. The UF6 gases were used as carriers of the uranium isotopes in the separation process and separation was effected by the difference in atomic mass of the two isotopes. The breakthrough of commercial gas separation membranes came much later, in 1980, with Monsanto’s Prism hollow fiber (HF) membrane made of polysulfone (PS) separating hydrogen from ammonia (Henis and Tripodi, 1980). This sparked the application of new membrane materials and in the mid-1980s companies like Cynara, Separex, and Grace Membrane Systems were using membrane plants for removing CO2 from methane in natural gas (NG), or desouring, a process close to the task of upgrading biogas to NG. These membranes, made of cellulose acetate (CA), paved the way for numerous new applications with a wide variety of different membrane materials. In the task of separating CO2 from methane, the first important development was the use of polyimide (PI) membranes in 1994 by MEDAL, an Air Liquide company (Baker and Lokhandwala, 2008). This was later improved by applying cross-linking and thermal rearrangement of the polymers in the membrane matrix (Chen et al., 2017); although an industrial application still must await maturation of this development. Other materials for CO2/CH4 separation are of the molecular sieving type, which employs ceramic zeolites or carbon nanomaterials known as carbon molecular sieves (CMSs). These materials have a very high degree of selectivity and permeability but suffer from high production costs and brittleness of the membrane. Therefore CMSs still need further development before they can achieve proper commercialization (He et

al., 2018). In Fig. 6.1 an updated timeline for the development of gas separation membranes is shown.

Figure 6.1 Timeline for the development of gas separation membranes.

Biogas production worldwide is growing linearly and the prospect for global production in 2022 is of 33 Mtoe/year (WBA, 2018), which would double the production over the last 10 years. This growth is mainly effected by an increase in production in Europe, but areas like East Asia are now contributing significantly. Furthermore, the incentive of using organic waste as feed for biogas production is influencing many other parts of the world (Kapoor et al., 2019; Raboni et al., 2015). This increase in production is followed by demand for the methane fraction of the gas mainly for transport fuel. In the communication from the European Commission to the EU parliament in 2011 regarding the Energy Roap 2050 (Commission, 2011), which is the guideline for a complete change of energy supplies from fossil to renewable sources, biogas is mentioned as one of the possible alternatives to fossil fuel, albeit not mentioning explicitly the need for upgrading biogas to NG quality. However, it is well known that an apprehensive use of biogas as a transport fuel or as fuel distributed in gas grids requires low-cost and reliable upgrading technologies, where membrane processes can indeed play a significant role. CO2 removal from biogas can be achieved in several ways including waterscrubbing, pressure swing adsorption, cryogenic separation, and membrane separation (Ullah Khan et al., 2017; Angelidaki et al., 2018). Among those, membrane technology stands out as an excellent opportunity regarding simplicity and flexibility, whereas it comprises low energy expenditure for the process associated with low costs. In membrane-based gas separation processes, the driving force for the separation is the difference in partial pressure of the gas species across a membrane barrier. The individual components can be separated attending to their different permeability rates through a porous membrane material, or to different solubilities and diffusivities through a nonporous, dense membrane material. In a status report made for the ICOM conference in 2002, the following advantages of membrane technology were highlighted (Drioli, 2002):

• Does not involve phase changes or chemical additives;

• Simple in concept and operation; • Modular and easy to scale up; • Greater efficiency for raw materials use and potential for recycling of byproducts; • Equipment size may be decreased.

In the comprehensive review “Membrane gas separation technologies for biogas upgrading” (Chen et al., 2015) and the report “Review of Biogas Upgrading Future Gas project, WP1” (Hjuler Klaus, 2017a) this is formulated more precisely: The currently most used methods for upgrading biogas to methane are compared with regard to energy consumption and economy, concluding that membrane technology is found to be at least as efficient as other methods, especially when utilizing the latest technology using cascade configurations giving very low off-gas losses. Membrane-based biogas upgrading has been known for the past few decades. However, research and development seeking novel membrane materials and process configurations is continuously required to improve the selectivity and productivity of these technologies. In the following sections, the principles of membrane gas separation processes, the development of new membrane materials, and the process configurations available are discussed with a special focus on biogas upgrading. The intention is in this way to provide a thorough explanation of the ways in which membrane technology can be applied for the task of upgrading biogas.

6.2 Basic of gas separation on membranes

Physically, gas separation in a membrane is performed on a mixture of gases fed to the membrane unit by subjecting the mixture to an elevated pressure, with flows as shown in Fig. 6.2.

Figure 6.2 Gas composed of feed stream FF with concentrations Ci is separated into two streams: the Retentate with flow FiR and concentration CiR, and the Permeate, with flow FiP and concentration CiP.

In this context, the basic equations that describe the material flow in the process are: Flow balance:

(6.1)

Mass balance for component i:

(6.2)

And for design of the membrane processes, other relevant parameters are: The ratio of feed pressure to permeate pressure:

(6.3)

And the ratio of permeate to feed flow (or stage-cut):

(6.4)

Effective separation takes place when some gas components through the membrane barrier quicker than others. This process can be described by three general transport mechanisms: Knudsen diffusion, solution-diffusion, or molecular sieving. The different mechanisms apply when the membrane satisfies specific structural conditions, as shown in Fig. 6.3.

Figure 6.3 Schematic porous and dense membranes and their gas permeation mechanisms. Adapted from Baker, R.W., 2004. Membrane Technology and Applications. John Wiley & Sons, Ltd, pp. 96–103.

The performance of the membrane for separation of a given binary gas mixture of components A and B is typically expressed as follows:

(6.5)

where D is the diffusivity coefficient and S the solubility constant for the given gases. The ratio (PA/PB) is often referred to as the permselectivity of the membrane and in the ideal case (ideal gas conditions) it is also referred to as the ideal membrane selectivity. In the nonideal case where several gases are present in the mixture, the permselectivity is better represented by:

(6.6)

where yA is the mole fraction of the gas component A in the permeate and xA is the mole fraction of the component A in the feed. Permeance is another parameter linked to the permeability, which is directly related to the thickness, l, of the membrane and as such is defined as pressurenormalized steady-state flux, P/l. The permeance will characterize the gas transport through the membrane. Permeance is an important parameter when comparing the separation suitability of membranes for mixed gases. A practical

unit often used is the GPU (gas permeation unit), as seen below:

(6.7)

In the SI system, permeance is expressed in the following unit:

(6.8)

However, another widely used and accepted unit for permeability is the Barrer:

(6.9)

Generally, membranes have been graded according to permeability and separation factor in a Robeson plot (Robeson, 2008). In this representation, the logarithm of αAB for a number of membrane materials and gas components to be separated into A and B is plotted against the logarithm of permeability for the faster of the two components. Using the plot, different membrane materials can be compared in of separation performance for a given gas pair. The Robeson plot reveals the known fact that a high separation factor is counteracted by a low permeability, and vice versa. In fact, the plot gives an indication of the trade-off between selectivity and permeability by defining an upper bound that limits the combined improvement that can be achieved in the separation of a given binary gas mixture (Freeman, 1999).

For porous membranes, effective gas separation requires that the mean free path of the molecules is larger than the average pore diameter. In that case, the gases experience Knudsen diffusion, where small molecules penetrate faster than heavier molecules. This effect is described by the Graham law of diffusion, which states that the transport rate of a gas is inversely proportional to the square root of its molecular mass. In some cases, the separation in a Knudsen flow situation is enhanced by the preferential adsorption of some components onto the membrane material. This results in diffusion of the molecules along the surface of the pores, and thus the selectivity of the separation process increases (Hägg and He, 2011). In the cases where some molecules are larger than the pore diameter, the separation takes place by molecular sieving. In this mechanism, the large molecules are excluded from age, whereas others of smaller size permeate following Knudsen flow and/or preferential adsorption diffusion flow. Porous membranes for gas separation remain a minor part of the commercial market, which is still dominated by dense membrane designs. Some processes based on porous membranes have gained attention in recent years, including CMSs, mixed-matrix membranes (MMMs), and thermally rearranged (TR) mixed-matrix membranes (Hamid and Jeong, 2018). A summary of the recent advancements in these techniques and their application to biogas upgrading is presented in Section 6.6. The most common type of gas separation membranes are dense membranes, consisting of thin films without defined pores. Dense membranes are characterized by the solution-diffusion mechanism, where the transport of gas molecules is governed by dissolution in the membrane material and diffusion through it. Dense polymer membranes for gas separation can be described as having four structural levels, each of which influences the performance of the separation process (Hoehn, 1985):

Level I: Chemical composition of the polymer that forms the selective membrane layer; Level II: Steric relationships in repeat units of the selective polymer;

Level III: Morphology of the membrane’s selective layer; Level IV: The overall membrane structure, including structural relationships between the selective layer and the rest of the membrane.

Levels I and II concern the chemistry of the polymer forming the membrane, and the factors influencing the diffusion of gas components through it. These will be described further in Section 6.3. Levels III and IV are related to the membrane symmetry, that is, whether the morphology is symmetric and even through the entire cross section of the membrane, or if it is asymmetric and built from different structures. Practically all materials used for fabrication of membranes for CO2 removal are polymer-based, for example, CA, PIs, polyamides, PS, polycarbonates, and polyetherimide. Table 6.1 shows the polymeric materials most commonly used for CO2 removal.

Table 6.1

Polymer

P (CO2) Barrer

α (CO2/CH4)

Reference

Cellulose acetate (CA)

8.9

20–25

Spillman (1995)

Polyimide (PI)

65–110

25–15

Ismail et al. (2015)

25–368

13.1–34.5

Álvarez et al. (2020)

Polyamide (PA)

—

12.8–14.4

Sridhar et al. (2007)

Polysulfone (PSf)

5.6

22.4

Sanders et al. (2013)

Polycarbonate (PC)

6.8

19

Koros and Fleming (1993)

Polyetherimide (PEI)

1.25

17.5

Vega et al. (2019)

Most commercial membranes are made as asymmetric membranes consisting of a thick porous layer or , on the top of which a thin membrane layer is placed. This layer governs the separation and is also referred to as the selective layer or skin layer. The selective layer can be either formed on the layer as an integral part of the whole membrane or incorporated as a coating. In the latter case, the membrane structure is called a composite membrane. Next, an additional level can be added concerning the way the membrane is geometrically constructed. Two different geometries are typically envisaged: HF or flat sheet. Hollow fibers are very thin polymer tubes (50–3000 μm diameter) that can be packed in compact modules and provide a high area-to-volume ratio. Typically, the fibers are composed of a porous that contains the selective layer either on the interior or exterior of the fiber. A thorough description of their design and manufacture is given in Dortmundt and Doshi (1999) and is sketched in Fig. 6.4. This last level also concerns the way the individual membranes are packaged into a complete membrane element. Hollow fibers are often packed in bundles and flat sheet membranes are often arranged in spiral wound elements (Dortmundt and Doshi, 1999), as shown in Figs. 6.5 and 6.6

Figure 6.4 Sketch of the build of a hollow fiber and flat sheet membranes where the selective layer and are clearly defined. Adapted from Dortmundt, D., Doshi, K., 1999. Recent developments in CO2 removal membrane technology. UOP LLC 1, with kind permission and courtesy of Honeywell UOP.

Figure 6.5 Sketch of a spiral-wound module. Adapted from Dortmundt, D., Doshi, K., 1999. Recent developments in CO2 removal membrane technology. UOP LLC 1, with kind permission and courtesy of Honeywell UOP.

Figure 6.6 Hollow-fiber membrane element. Adapted from Dortmundt, D., Doshi, K., 1999. Recent developments in CO2 removal membrane technology. UOP LLC 1, with kind permission and courtesy of Honeywell UOP.

In the spiral-wound arrangement, two flat sheets of membrane with a permeate spacer in between are glued along three of their sides to form an envelope (or leaf, as it is called in the membrane industry) that is open at one end. Many of these envelopes are separated by feed spacers and wrapped around a permeate tube with their open end facing the permeate tube. The feed gas enters along the side of the membrane and es through the feed spacers that separate the envelopes. As the gas travels between the envelopes, CO2, H2S, and other highly permeable compounds permeate into the envelope. These permeated components have only one outlet: they must travel within the envelope to the permeate tube. The driving force for transport is the lowpermeate and high-feed pressures. The permeate gas enters the permeate tube through holes drilled in the tube. From there, it travels down the tube to the permeate from other tubes (Dortmundt and Doshi, 1999). Any gas on the feed side that does not get a chance to permeate leaves through the side of the element opposite the feed position (Dortmundt and Doshi, 1999). In hollow-fiber elements, very fine HFs are wrapped around a central tube in a highly dense pattern. In this wrapping pattern, both open ends of the fiber end up at a permeate pot on one side of the element. Feed gas flows over and between the fibers, and some components permeate into them. The permeated gas then travels within the fibers until it reaches the permeate pot, where it mixes with the permeates from other fibers. The total permeate exits the element through a permeate pipe as shown in Fig. 6.6. In gas–liquid membrane ors, the separation takes place via the selective absorption of some of the gas components into a liquid phase. Microporous hydrophobic membranes, often made as HFs from polypropylene, separate the gaseous from the liquid phase. As shown in Fig. 6.7, molecules from the gas stream, flowing in one direction, are able to diffuse through the membrane, and are absorbed on the other side by liquid flowing in counter current mode. The

liquid is prevented from flowing to the gas side due to slight pressurization of the gas. These membranes work at approximately atmospheric pressure (100 kPa), which allows low-cost construction, and they have a very high selectivity.

Figure 6.7 Membrane gas absorption principle. Adapted from Ismail, A.F., Khulbe, K.C., Matsuura, T., 2015. Gas Separation Membranes. Switzerland: Springer.

Employing membranes rather than packed columns for gas–liquid ing and absorption offers several advantages like independence from flooding, weeping, and over-loading, hence giving a better freedom of choice of masstransfer coefficient and area that can be varied independently. Several reported experiments with this technology have been published (Delgado et al., 2009; Yan et al., 2014). As reviewed by Zhang and Wang (2013), large-scale application of gas–liquid membrane ors for CO2 absorption would require further advancements in the following topics:

1. Improvement of present microporous membranes in of antiwettability, and thermal and chemical resistance; 2. Development of novel absorbents with high separation effectivity, good compatibility with the membrane, and with the possibility of cost-efficient regeneration; 3. Development of novel membrane configurations and modules that enhance mass transfer.

Gas–liquid absorption membranes for biogas upgrading have been developed only recently and are still in their emerging phase. Several investigations have reported CO2 removal from gas mixtures using adsorbents like demineralized water (Heile et al., 2014), amino acid solutions (Yan et al., 2014), amines like monoethanolamine or diethylamine (Nakhjiri et al., 2018), and ionic liquids (Estahbanati et al., 2017). The removal of CO2 carried out with a potassium Largininate solution is very efficient; biogas with 55% CH4 can be upgraded to more than 99% CH4 in one step (Yan et al., 2014). The solution can subsequently be regenerated by heating, which releases a pure CO2 flow that can be sold for industrial applications.

Nevertheless, most of the present membrane-based biogas upgrading plants use hollow-fiber composite membranes consisting of a dense selective layer placed on top of a porous . This type of membrane is typically based on selective layers of PI or CA, that provide good CO2/CH4 selectivity, but require high operating pressures up to 20 bar (Park et al., 2017). There has been growing interest in shifting toward the integration of membrane separation and absorption in membrane ors that combine the advantages of both unit operations.

6.3 Membrane materials and structures

6.3.1 Polymer structures and their influence in permeation

As introduced earlier, permeability results from the mobility of gas molecules through the membrane material (diffusion) and their solubility inside the membrane material (dissolution), whereas selectivity s for preferential age of some components over others. Polymer membrane materials are typically amorphous and thus behave very differently whether they are above or below the polymer glass transition temperature (Tg). Below Tg, the polymer chains are rigid, and the material is referred to as a glassy polymer. Glassy polymers have mobility selectivity, meaning that the selectivity is directly related to the difference in sizes of the gas species and how fast they diffuse through the membrane material. In general, glassy polymers have lower permeation rates, and favor the age of smaller molecules. Above Tg, a certain degree of rotation and flexibility around the polymer backbone is allowed, converting the polymeric membrane into a rubbery material. Rubbery polymers have sorption selectivity, which is mainly influenced by the differences in condensability and sorption of the different gas species in the membrane. Rubbery membranes have typically larger permeation rates, and favor more condensable molecules, that are typically larger. Fig. 6.8 illustrates the typical relations between diffusion and permeant size (1), and diffusion and condensability (2). For a given gas pair a and b, diffusion selectivity would favor the age of a over b, whereas the higher condensability of b would preferentially allow age of b over a (Drioli and Giorno, 2009).

Figure 6.8 Representation of variation of the diffusion coefficient with permeant size (1) and solubility coefficient with permeant condensability (2) for rubbery and glassy polymers.

A very thorough review on polymeric membranes has been made by Yampolskii (2012) in which it is stated that although polymeric membranes can be used today for separation of nearly all imaginable gas mixtures, large industrial applications have been realized only for the following processes (Baker and Lokhandwala, 2008; Baker, 2002; Bernardo et al., 2009; Bernardo and Drioli, 2010; Ockwig and Nenoff, 2007):

• Air separation (production of technical-grade nitrogen-or oxygen-enriched air); • Hydrogen separations (separation of the mixtures H2/N2, H2/CH4, and H2/CO in refineries and petrochemical industry); • Separation of CO2/CH4 mixtures (NG “sweetening,” upgrading biogas) and CO2/N2 (treatment of flue gas, etc.); • Vapor/gas separations (numerous systems).

Table 6.2 presents some of the leading companies that are active in industrial membrane gas separation and the principal membrane materials used.

Table 6.2

Company

Membrane materials used

Permea (air products)

Polysulfone

MEDAL (air liquid)

Polyimides

Generon

Tetrabromopolycarbonate

Separex (UOP, now Honeywell UOP)

Cellulose acetate

Aquila

Poly(phenylene oxide)

Ube

Polyimide

MTR

Silicon rubber

Helmholtz Centrum (formerly GKSS)

Silicon rubber

Grasysa

Polyimide, polysulfone

Kryogenmasha

Poly(vinyltrimethylsilane), tetrabromopolycarbonate

Air Liquid

Ethyl cellulose

OPW Vaposaver

Poly(trimethylsilyl propyne)

aProducer of membrane modules and installations only.

Most of these polymeric membranes were designed in the mid-20th century and comprised polymers like polyolefins, vinylic polymers like poly (vinyl acetate), polytetrafluoroethylene, siloxane, and other rubbers, and cellulose derivatives. Studies on these polymers revealed that there exist straightforward correlations between the chemical structure of the polymer repeat unit and the observed gas permeation parameters like the diffusivity and condensability of the gases. The most detailed results were obtained for siloxane polymers with different side groups attached to the Si atom. It was shown that when the size of the side groups increased, the flexibility of the polymer chains decreased. This resulted in a higher glass transition temperature and lower permeability, that is, changing the side group in a siloxane from a methyl (CH3) to a phenyl group (C6H5) reduced the permeability by 95% (Stern et al., 1987). Another important group of polymers from that period is the glassy polymers, where the gas permeation properties are closely related to the polymer structure of the repeated side chains, which is formulated generally for numerous Sisubstituted glassy polymers for enhancement of permeability (Yampolskii, 2012; Nagai et al., 2001; Yampolskii, 2007). The main structural characteristics and their effects are the following:

1. Symmetry of side groups. The strongest effects are observed for completely symmetrical substituents; for example, the effect of the SiMe3 group is stronger than that of SiMe2Et. 2. A bulky group should be attached directly to the main chain and not via a spacer. 3. Long linear substituents (e.g., SiMe2-n-C6H13) show significant decreases in permeability.

Explanations of these phenomena can be given based on the free volume theory and the application of probe methods, as described excellently in the book by Ismail et al. (2015). The free volume of a polymer membrane is in general the free space among the polymer chains that is not filled with the atoms of the polymer. In a rubbery polymer, the free volume may vary as the polymer chains are able to shift position, enabling a higher diffusion of permeant molecules, whereas in a glassy polymer, the polymer chains are locked in position relative to one another, which often leads to a low diffusion of the permeant molecules. This again leads to a “trade-off” between selectivity and permeability, that is, polymers with high permeability have low selectivity, and vice versa. This has been given a very comprehensive treatment in the paper by Alentiev and Yampolskii (2000). In polymers with aromatic backbones (PIs, polycarbonates, PSs), a marked increase in permeability can be achieved by symmetrical or asymmetrical substitution of phenylene rings (Pixton and Paul, 1994). Different radicals were used in such substitutions, for example, and other linear and branched alkyls, CF3, Cl, and so on. Examples of the effects of such modification of diamine moiety of 6FDA PIs are given in Table 6.3 (Yampolskii, 2012).

Table 6.3

R1

R2 H

P(O2), Barrer H

2.8

α (O2/N2) 5.65

Substituent volume, Å³ 16.5

H

CF3

5.2

5.71

79.0

Me

Me

11.0

4.17

90.8

Et

Et

18.4

4.20

158.8

i-Pr

i-Pr

47.1

3.76

226.8

The increase in permeability is correlated with the total size of side groups as shown in the last column of Table 6.3. These data indicate that substitution results in a looser chain packing or growth of free volume, which is accompanied by decreases in permselectivity α(O2/N2). Another element of the design of polymers with aromatic backbones is the structure of connector groups. These effects are also revealed in the transport properties of polycondensation materials (Pixton and Paul, 1994). Bulkier connector groups can stiffen the chain backbones, inhibit their dense packing, and thus increase free volume and permeability. This situation is illustrated for the family of PIs in Table 6.4.

Table 6.4

Connector group X

P(CO2), Barrer

α (CO2/CH4)

–O–

23

60.5

–CH2–

19.3

44.9

–C(CH3)2–

30

42.9

–C(CF3)2–

63.9

39.9

Adapted from Coleman, M., Kohn, R., Koros, W., 1993. Gas-separation applications of miscible blends of isomeric polyimides. J. Appl. Polym. Sci. 50(6), 1059–1064.

A modification of dense polymeric membranes termed TR polymer membranes is carried out by applying heat to solid glassy polymers like PI, and the resulting deformation transforms the molecular chains and produces micropores within the polymers, increasing the permeability of gas molecules. Such polymers, when used in membranes, have the potential to augment selective permeability (i.e., high permeability and selectivity) for the target gas by acting as a molecular sieve, and the membranes have a large free volume and an enhanced masstransfer rate due to micropores within the polymers themselves. Even though they provide exceptional results on a laboratory scale, TR carbonized membranes are often brittle and hence suffer from mechanical problems when large-scale modules are envisioned (Hägg and He, 2011). However, in the review by Jeon and Lee (2015) the conclusion is that TR polymers have a fair chance of exceeding the Robeson upper bound, which is a graphical correlation of selectivity and permeability for different membrane types and is described in more detail in Section 6.4, and therefore appear as promising candidates for future use in biogas upgrading.

6.3.2 Inorganic membranes for gas separation

Gas separation inorganic membranes are mainly divided into ceramic membranes and zeolites. Ceramic membranes are produced by coating of ceramic layers on top of a and are typically formed of aluminum, titanium, zirconium or silicon oxides, as well as silicon carbide. Zeolites are three-dimensional networks of silicate and aluminosilicate that can

contain pores in the range of 3–8 Å (Baker, 2004). Both can withstand high temperatures and have good chemical resistance. However, they are expensive and brittle. Zeolites have been used dispersed in polymers in the form of mixedmatrix membranes (Pechar et al., 2006; Bastani et al., 2013) or as for gas–liquid membrane ors (Zhang et al., 2018).

6.3.3 Carbon molecular sieve membranes

In a category of its own between TR polymers and inorganic membranes, it is possible to allocate CMSs, which when made from CA are a relatively cheap eco-friendly material well suited for the application of CO2 removal from biogas. CMSs are most often produced from HFs made from a precursor like CA, enabling the production of membranes with close packing and a large area. Materials chosen as precursors should not be the cause of any defects in the finished carbon structure forming the membrane (Saufi and Ismail, 2004), therefore they must preferentially be thermosetting polymers (polymers that do not flow before decomposing), such as PIs, poly-furfuryl alcohols, phenolic resins, polyvinylidene chloride, polyacronitrile, or cellulose derivatives like CA. These choices are highlighted in the review of CMS manufacture and applications by Hägg and He (2011), where it is stated that producing a highperformance carbon membrane is a quite difficult task, since it includes many steps that must be well controlled and optimized. The fabrication process for CMS membranes normally consists of six important steps, that is, raw material selection, material functionalization (additives for given optimal function), precursor preparation, pretreatment, carbonization (pyrolyzation), and posttreatment, as illustrated in Fig. 6.9.

Figure 6.9 Important steps in the manufacture of CMS membranes. Adapted from Hägg, M.-B., He X., 2011. Chapter 15. Carbon Molecular Sieve Membranes for Gas Separation. 162–191.

Even when complying with a strict protocol in the manufacturing process, it is for some polymer precursors often encountered that the CMS membranes produced are brittle and mechanically feeble, making packing in dense modules with a high area a difficult task (Hägg and He, 2011; Lagorsse, 2004). However, it has been shown in many applications, where cellulose, CA, or cuprammonium cellulose have been used as precursors, that the resulting CMS fibers were both flexible and with high mechanical strength (Lagorsse, 2004; Soffer et al., 1999). Carbonization is performed in an inert atmosphere through a protocol in which heating up to 800°C and cooling is performed in strictly controlled timing sequences and the addition of different additives dependent on the precursor polymer chosen (Hägg and He, 2011; Prangsgaard Andersen, 2016) (Fig. 6.10).

Figure 6.10 Protocol for the carbonization of a precursor to make a carbon molecular sieve. Adapted from Prangsgaard Andersen, T., 2016. Hollow Fiber Carbon Membranes for Separation of Hydrogen and Carbon Dioxide (PhD thesis), University of Southern Denmark.

The Israeli company Carbon Membranes Ltd. made full-scale CMS membrane modules in 1995–2001 with both excellent mechanical properties and good selectivity and permeability for separating carbon dioxide from methane based on cellulose fibers as precursors. The process is described in detail in the correspondent patent (Soffer et al., 1999). This process was further used for production of CMS membranes with excellent results for the separation of hydrogen and carbon dioxide in a PhD study at the University of Southern Denmark (Prangsgaard Andersen, 2016).

6.3.4 Mixed-matrix membranes

In order to solve the problems of high cost, difficulty to fabricate, and brittleness associated with zeolites and CMSs, MMMs were introduced. These new membranes consisted of a polymeric matrix where particles of highly selective materials such as zeolites or CMSs were dispersed. This resulted in hybrid media that combined the high selectivity observed for the dispersed particles with the desired processability and economy of polymeric matrix (Moore et al., 2004). As reviewed by Ozturk and Demirciyeva (2013) many materials have been investigated in MMMs for improved CO2/CH4 separation ranging from zeolites to CMSs, metal oxides, and metal-organic frameworks (Ozturk and Demirciyeva, 2013). Vu et al. (2003) reported increased CO2 permeability and CO2/CH4 permselectivity for PI and polyetherimide membranes loaded with CMSs. The CO2 permeability increased from 10.0 to 12.6 Barrer for the original and 36 vol.% CMS-loaded PI; and from 1.45 to 4.48 for the original and 35 vol.%

CMS-loaded polyetherimide. The CO2/CH4 permselectivity also increased, namely from 38.8 to 53.7 for the original and 36 vol.% CMS-loaded PI, and from 38.8 to 53.7 for the original and 35 vol.% CMS-loaded polyetherimide. For orientation, the CMSs used had a CO2 permeability of 44 Barrer and a CO2/CH4 permselectivity of 200. In 2015, Mahmoudi et al. (2015) synthesized and characterized MMMs composed of three phases: a polymer (PEBA), a liquid (PEG), and a zeolite (nanozeolite X). Incorporation of PEG resulted in an increase of ethyl oxide units, that resulted in enhanced CO2 dissolution and permeability. These membranes provided excellent permeability and CO2/CH4 selectivity above the Robeson’s upper bound. However, commercialization of MMMs is not yet a reality, due to the difficulty of their production and their easy fouling due to absorption of impurities (Baker, 2004). Some commercial rubbery materials with high affinity for CO2 separation in MMMs are currently available. One material that stands out as a suitable candidate for CO2/CH4 separation is the thermoplastic elastomer PEBA, a family of poly(ether-block-amide) copolymers that are commercialized under the tradename PEBAX (Esposito et al., 2015). These copolymers combine the rigidity provided by the amide groups, with the elasticity and CO2 affinity given by the ethylene oxide groups (Wahab and Sunarti, 2015). Unfortunately, PEBAX composite hollow-fiber membranes cannot be produced by direct spinning without resulting in unstable or defect fibers. Therefore PEBAX are nowadays preferred as in MMMs or as a selective layer on composite membranes (Esposito et al., 2015).

6.3.5 Results of membrane operations with different materials

Of the above-mentioned membrane materials, PI membranes have been tested for upgrading biogas in a test rig with an UBE HF module A-2 (Harasimowicz et al., 2007). The result was that a gas with 50% CH4/50% CO2 could be upgraded in one up to a methane concentration in the retentate of 93%, but with a loss

of no less than 12% of methane in the permeate. In a once-through set up with PI HF elements from UBE and polysulfone HF elements from Air Products arranged in parallel for comparison, a methane concentration of over 95% was achieved. However, the retentate contained over 25% methane, giving a severe loss (Vrbová and Ciahotný, 2017). This can, however, be improved by changing to a two-stage or three-stage process with appropriate recycling of the permeate stream as described in the review by Haider et al. (2018). An interesting aspect is described in a paper by Makaruk et al. (2013) in which rubbery membranes [poly-(di-methyl siloxane)] are used for separating hydrogen sulfide found in the biogas together with CO2 from methane and the retentate from this filtration is fed to a glassy PI membrane, where the residual carbon dioxide is separated from methane obtaining a high-purity methane flow without hydrogen sulfide. A very extensive review of different polymers, MMM and PEBAX membranes used for carbon capture including biogas upgrading has been given by He (2018). CMSs are described with regards to the production process in an excellent description by Haider et al. of CMS HFs (Haider et al., 2018). CMS HF fullmembrane modules with areas around 5 m² have been tested in large pilot plants also for biogas upgrading purposes (Norddahl and du Preez, 2007).

6.4 Theory of transport in gas separation on membranes

6.4.1 Transport through rubbery polymers

Rubbery polymers act in such a way that gases get absorbed in the membrane material following Henry’s law:

(6.10)

where [Ci]liq is the concentration of gas component i dissolved in the polymer. Hi is the Henry’s coefficient for component i, and pi is the partial pressure in the gas phase of component i. Relations between polymer structure and permeation rates of gases in the polymer are described in detail by Matteucci et al. in the book Materials Science of Membranes for Gas and Vapor Separation (Matteucci et al., 2006). This treatise is based on the solution-diffusion model for permeation of the gases in the membrane, where the permeants from one side of the membrane to the other side driven by a concentration gradient. In the case where the permeation is taking place in a symmetric flat membrane of the thickness l as shown in Fig. 6.11 the following relation applies:

(6.11)

where φ is the flux or steady-state permeation rate, P is the permeability coefficient, Δp=P2 − P1 is the pressure difference between the two sides of the membrane, and l is the thickness of the membrane.

Figure 6.11 Representation of gas or vapor transport through a nonporous polymeric membrane of thickness l. P2 will be greater than P1

In a fundamental formulation, the basis for permeation is phrased by Fick’s first law at steady state at any point inside the membrane:

(6.12)

where φ is the flux, Dloc is the local diffusion coefficient of the gas in the polymer, w is the mass fraction of the gas in the polymer, and x is the distance perpendicular to the faces of the flat membrane. The flux for a component i in a given gas mixture (φi) can thus be expressed in a simple form as:

(6.13)

where Di is the average diffusion coefficient in the polymer, Hi is a common denomination for the Henry’s sorption constant or coefficient, but in order to comply with the nomenclature often used in the gas separation literature we shall use the designation Si for this term, and Δpi is the difference in partial pressures of component i between the two planes of the polymer skin layer. Generally, the fugacity of the gas components should be used in these relations. However, gas separation is normally performed under conditions of relatively low pressures and temperatures, near ideal conditions. Eventually this led to an expression of the permeability as:

(6.14)

P, D, and S values in polymers can be determined using different strategies and these methods have been thoroughly reviewed by Felder and Huvard (1980) and Lin and Freeman (2005). In membrane gas separation the permeability coefficient P is ideally close to a constant for the material, relatively independent of the composition and pressure of the feed and permeate gases, as explained in the thorough review and in-depth discussion of this model presented by Wijmans and Baker (1995). However, this statement should be treated with some caution. For example, the permeability of vapors at partial pressures close to saturation often increases substantially with increasing partial pressure. This effect is commonly ascribed to plasticization and occurs when the diffusion and activity coefficient of the permeant gases change due to changes in the polymer matrix under high permeant partial pressures (Wijmans and Baker, 1995). Improvements of this condition are treated further in the discussion of modified membrane structures below. The properties of the gas will also be important for the gas separation: an inert gas (like H2, N2, also CH4) has very low critical temperature and will not easily dissolve in the material. Hydrocarbons, however, and nonideal gases like CO2, have relatively high critical temperatures and will easily dissolve in the material and might also swell the membrane. This will naturally affect the separation, and the selectivity will go down. It is thus of major importance to evaluate the suitability of the material for the separation of the gas pair in question and with respect to process conditions for the potential application.

6.4.2 Transport equations through glassy polymers

For glassy polymers, an equilibrium state cannot be achieved because of the

limited mobility of the polymer chains below Tg. The final state is determined by the history of the polymer state before transition to glassy state. Sorption in glassy polymers can be explained with the dual-mode-sorption (DMS) model described in Koros et al. (1977) and Koros and Paul (1981). In this model, some components act as if sorbed in a low-molecular-weight liquid or rubbery polymer, as described by Henry’s law relation. The other population of sorbed material is viewed as being due to uptake into the unrelaxed volume or “microvoids” present in the glassy polymer. This population is described as a Langmuir “hole-filling” process. The total sorption is the sum of these two mechanisms, and is expressed as follows:

(6.15)

where Ki is the Henry’s law constant, pi is the penetrant pressure, [Ci]L is the Langmuir capacity constant, and b is the affinity constant. The component parts that make up the DMS model are illustrated in Fig. 6.12 (2) as a concave curve.

Figure 6.12 Typical gas sorption isotherm form for polymeric media. Adapted from Koros and Chern (1987).

Simple homogeneous swelling of rubbery polymers by compatible penetrants tends to produce convex isotherms contrary to Fig. 6.12 (2). More complicated isotherms involving an inflection are observed for vapors and even highly sorbing gases such as CO2 in glassy polymers under certain conditions. The DMS model suggests that two different environments co-exist in equilibrium in glassy polymers due to rapid exchange between the two environments. In one environment the penetrant is dissolved in that particular media, whereas in the adjacent environment the penetrant component “jumps” across the adjacent “microvoid” spaces. For present purposes, a simple polynomial fit of C versus p to any of the data forms shown in Fig. 6.12 would suffice to evaluate dC/dp for use in determination of the solubility constant S and the diffusivity coefficient D. Such an approach is termed phenomenological in the sense that it simply describes the phenomenon in question, without considering its physical bases. Methods for obtaining relevant data are given in the book Gas Separation Membranes: Polymeric and Inorganic by Ismail et al. (2015). While the DMS model allows to obtain good correlations and useful insights into the polymeric structure when specific experimental data are available, it cannot be used in a predictive mode to evaluate gas or vapor apparent solubility in glassy polymer phases, as stated in the paper by Doghieri et al. (2006). Models such as nonequilibrium thermodynamics and statistical-associating-fluid theory are based on reliable equations-of-state of polymers and will probably eventually replace the conventional DMS model. An important achievement of these more advanced models is their ability to predict sorption isotherms based on thermodynamic properties, a quality the DSM model does not provide. These nonequilibrium models are well described in Chapter 6 of the book Materials Science of Membranes for Gas and Vapor Separation (Wiley, 2006) (Freeman et al., 2006) with numerous references. In a recent publication using the nonequilibrium lattice fluid model on several glassy polymers, experimental results were well in accordance with theory. The thermodynamic model describes penetrant permeability in glassy polymers and the model represents

experimental behaviors with and without plasticization (Minelli and Sarti, 2013). A Robeson plot for the separation of the binary mixture CO2/CH4 is shown in Fig. 6.13 (Robeson, 2008; Park et al., 2017; Park et al., 2007 where the upper bound is clearly defined.

Figure 6.13 Robeson plot and upper bound correlation for CO2/CH4 separation [filled squares are TR polymers with different pretreatments as described in Park et al. (2007)].

In the case of separation of CO2 and CH4, the upper bound is evidently raised to a new level by the exceptional separation properties of the so-called TR polymers, that were described in an earlier section in this chapter. However, none of these polymer types is presently utilized at an industrial scale. Currently, polymer membranes for gas separation are chosen as close to high permeance and selectivity as possible. This, however, induces concentration polarization, a phenomenon that might significantly limit the performance of the separation by increasing the driving force of the retained gas and reducing the driving force of the fast permeating species. This is especially observed in membranes with high selectivity (Mourgues and Sanchez, 2005). In the review by Jeon and Lee (2015), several polymer membranes in actual service for separation of CO2 and CH4, as well as other polymers still in the research phase, have been described and compared using the Robeson plot. The conclusion in this chapter is that polymers of the PI type [i.e., PI 6-FDATMODA/DAT (3/1)] have good permeability and selectivity. These values were observed close to or even exceeding the Robeson limit, which makes this type of polymer very suitable for future application for biogas upgrading. Compared to other polymers like PS and polycarbonate (PC), PIs have the further advantage that they are more chemically and physically stable in the environments where biogas upgrading takes place. In some conditions such as in high operating pressures or due to membrane aging, glassy polymer membranes may suffer from plasticization, where the sorption of some components results in increased polymer chain mobility and consequent selectivity loss. This process is sometimes also referred to as membrane swelling. Plasticization is an especially relevant issue in biogas upgrading applications since CO2 frequently acts as a plasticizer. When this happens, the permeability of methane increases and the separation process losses its reliability. Different methodologies have been developed to suppress or minimize the effect of plasticization in biogas upgrading. Some examples

include thermal treatment or cross-linking of the polymer in order to increase rigidity and reduce chain mobility. In a review by Baker and Low (2014) examples of modification of glassy PI membranes incorporating thermal rearrangements and cross-linking that improved the reduction in permeability and separation in the PI membranes used for CO2 separation from methane are reported. Bos et al. (1998) reported suppression of CO2-induced plasticization of PI membranes stabilized by thermal treatment and cross-linking at 350°C. Chen et al. (2012) removed plasticization in PIs by cross-linking with a diamino organosilicone [bis(3-aminopropyl)-tetramethyldisiloxane or APTMDS] methanol solution at ambient temperature. Summarizing, dense polymeric materials may thus be in their glassy or rubbery state depending on their glass temperature and operating conditions. Gas separation in dense membranes is basically described according to solution/diffusion. In glassy state, the diffusion of the gas component will determine the selectivity between the gases, while in rubbery state, the solution of the gas in the membrane material will be of major importance. Generally, the selectivity will be high and permeation low in a glassy, dense membrane, while selectivity will be low and permeation high in a rubbery material.

6.5 Membrane configurations and plant design for upgrading biogas

Design of units for separating methane from carbon dioxide is often taken from knowledge of purifying raw NG from impurities including CO2. Raw NG does, however, often contain higher alkanes like ethane, propane, and butane in relatively high concentrations and these gases often introduce complexities when using membrane for the separation, like plasticizing and loss of permeability of the membrane (Baker and Low, 2014). These components are, however, not found in biogas and impurities like hydrogen sulfide, ammonia, and water are relatively easy to remove from the raw biogas, as described below. Already, in 1983, Schell stated some conditions that are still relevant today (Schell, 1983):

• Flow is low to moderate: 5000–5,000,000 m³ (STP)/d—the average biogas plant production is now around 15,000 m³ (STP)/d and up to 75,000 m³ (STP)/d. • The feed should contain moderate concentrations (10%–65%) of the most permeable gas giving a reasonable partial pressure of the most permeable gas on the feed side, which is often the case with biogas. • The feed should preferentially have a moderate to high pressure [10–50 bar (g)] on the feed side to reduce cost of operation—but this is not normal in biogas operations. • For only membrane operations the product should not have a high purity requirement as that is not easily achieved in a membrane operation.

In the paper by Makaruk et al. (2010) describing membrane technology for upgrading biogas to NG quality, it is stated that the membrane process does not compete well with absorption and adsorption techniques, when methane recovery is concerned. This is true unless the high CO2-containing permeate

fraction is utilized for other purposes, that is, for heating the biogas digesters. Other uses are, however, also possible as described below. In this investigation, it is also concluded that in a pure membrane process for upgrading biogas, the concern about loss of methane in the final permeate implies more compression in a staged process, increasing the energy consumption and the overall cost of the process. In the paper by Makaruk et al. it is concluded that an optimal design could comprise a two-stage membrane configuration combined with a permeate burner supplying heat to the biogas digesters as shown in Fig. 6.14. In this configuration, the energy consumption is limited to about 0.3 kWh/m³ biogas treated.

Figure 6.14 Simplified flowsheet for the two-stage configuration with a permeate burner to produce NG substitute and heat.

Peppers et al. (2019) recently reported biogas upgrading from anaerobic digestion of food waste using PI hollow-fiber modules from Air Liquide in a two-stage configuration. The resulting biomethane did not meet the pipeline specifications in the region (California, United States) due to high CO2 remaining content. However, trace contaminants were efficiently removed by this process. On the other hand, Žák et al. (2018) reported a two-stage biogas upgrading process using dense hollow-fiber membranes made of polyester carbonate. The breakthrough of this study was the use of membrane modules resistant to the presence of water and H2S. Due to this implementation, the pretreatment for desulfurization and drying could be avoided, and the upgrading was achieved as a single-step method. Lower energy consumption combined with high CO2/CH4 selectivity resulted in 96% vol/vol CH4 purity and reduced capital expenditures when compared to other methods. A comparison between two-stage and three-stage membrane processes is depicted in Fig. 6.15.

Figure 6.15 Two-and three-stage membrane cascades for biogas upgrading (Makaruk et al., 2010).

6.6 Recent developments in membrane-based CO2/CH4 separation

Nowadays, large-scale membrane gas separation is mainly performed by polymeric membranes. Furthermore, biogas plant designs favor large plants where the feed material is collected from many sources in a large surrounding area. It is now common to see biogas plants in Europe fed with between 1000 and 2500 tons of organic material per day, which results in biogas production of the order of 35,000–75,000 N m³ of biogas/day or up to around 3000 m³ biogas/h with a methane content of 65 vol.% on average. The development searched for is sustainable use of the membranes at disposal, where a high purity of methane is achieved. At the same time, it is just as important that the CO2 separated from the gas stream finds an appropriate use. With a proper purification technology, biogas can be upgraded as fuel to be used in stoves/boilers, for producing electricity, for vehicles, for fuel cells, or it can be injected in the NG grid or used as a raw material for production of other commodity chemicals increasing the price of the methane. The recovered carbon dioxide has the potential to be used in enhanced oil recovery, carbonation of fly ashes (Mazzella et al., 2016) and wastes from aluminum extraction (Sun et al., 2015), and in food products. Traditionally, absorption processes like amine and water scrubbing and pressure swing absorption are the technologies used for upgrading biogas. These technologies have been inherited from NG cleaning or “de-souring” of NG, where carbon dioxide and hydrogen sulfide are removed from NG. In biogas from sources other than landfills (i.e., mixtures of manure and organic food wastes), light hydrocarbons are not present and the hydrogen sulfide can be reduced to a very low level of around 10 ppm by rather cheap micro-biological technologies (Guerrero et al., 2015), while at the same time, though, introducing nitrogen in the gas stream, which in turn might give problems with regard to injection in the NG grid. All the same, this gives membrane technology a competitive edge over absorption technologies both with regard to cost but also sustainability, as membrane technology requires less energy and supplementary

materials consumption than absorption technologies. The selection of the best upgrading technology is highly dependent on the biogas final utilization and different alternatives are available at commercial or demonstration plant levels. Here cryogenic separation, absorption, and their correspondent membrane-assisted hybrid separations were considered for their potential in reaching industrial implementation or for their possible optimization.

6.6.1 Biogas upgrading by cryogenic and hybrid cryogenic-membrane separation

The use of cryogenic separation is made possible by the difference in the condensing temperature between methane and carbon dioxide. It is performed by cooling the biogas in sequence with interstage cooling. However, due to the nature of the process some technical limitations need to be carefully considered. Water and other impurities must be preseparated in order to avoid the risk of icing, and carbon dioxide freeze-out must be also checked through examination of the temperature profile along all the process equipment. The necessity to use a refrigeration cycle fires up a warning regarding the energy consumption and the extent of the operative costs. Opinions regarding this technology are scattered. Baccioli et al. (2018) compared two small-scale bioliquefied NG plants. The former includes the classical upgrading technologies like water scrubbing plus regeneration, physical absorption, chemical absorption, pressure swing adsorption, and membrane technology. The latter is based on cryogenic separation. The results obtained for 97 vol.% purity of methane, showed how cryogenic separation can be competitive when compared to traditional upgrading systems. Hjuler and Aryal (Hjuler Klaus, 2017b) reported high energy consumption and the absence of large-scale industrial applications. The same conclusion was reached by Maqsood et al. (2014), highlighting also how this technology could represent a valid alternative among different upgrading options. In order to mitigate the disadvantage of the high energy requirement keeping the benefit of the high separation performance, a valid solution is to consider

cryogenic separation together with another unit operation. Moving from single separation methods to hybrid flowsheets it is possible to combine different unit operations to overcome their individual limitations (Errico, 2017). In this perspective, membrane separation has reached promising results in the separation of carbon dioxide and it appears to be the perfect candidate for the synthesis of hybrid flowsheets together with cryogenic separation. Brunetti et al. (2010) evaluated the possibility of using membrane separation for carbon dioxide removal from power plant emissions while other researchers focus their attention on membrane material development to improve the separation efficiency (Farashi et al., 2019; Ahmad et al., 2018). A possible hybrid membrane-cryogenic separation flowsheet was proposed by Song et al. (2018a) and is schematically represented in Fig. 6.16.

Figure 6.16 Hybrid cryogenic-membrane process. Reproduced from Song C., et al., 2018a. Efficient biogas upgrading by a novel membrane-cryogenic hybrid process: experiment and simulation study. J. Membr. Sci. 565, 194– 202.

In their set up, the raw biogas is initially compressed, cooled, and concentrated in a membrane unit. Operating the membrane at low temperature enhances the CO2/CH4 selectivity (Liu et al., 2013). The membrane permeate is concentrated in carbon dioxide that is compressed and chilled to the liquefaction temperature. The membrane retentate is constituted by biomethane that is liquefied to facilitate its transportation. The authors considered the recovery of the cold energy of the bio-LNG to avoid the use of external refrigeration and tested the system for polyamides and PS membranes. The results showed that PS membranes offered the best performances, reaching a purity of methane of 98 wt.% with a recovery of 98.2% and a global energy consumption of 0.83 MJ/kg CH4. It is worth mentioning that studies on hybrid cryogenicmembrane processes for biogas separation are very limited. However, this technology is available for CO2/N2 separation (Song et al., 2017, 2018b) or O2/N2 (Burdyny and Struchtrup, 2010), giving the foundation for an extension to biogas upgrading. An important contribution was also given by Liao et al. (2018) defining a methodology to design optimal hybrid cryogenic flashmembrane systems.

6.6.2 Biogas upgrading by absorption and hybrid absorption-membrane

Absorption is a very well-known procedure for carbon dioxide removal that can be used for biogas upgrading. The process consists of two main units, the absorber and the stripper. In the absorber the biogas flows in countercurrent with respect to an amine-based solvent. The solvent binds the carbon dioxide and is transferred to the stripping for its regeneration, while the gas enriched in

methane is recovered at the top of the absorber. Different works have been published on the topic, mainly focused on the modeling (Madeddu et al., 2018; Aronu et al., 2011) or solvent selection (Freeman, 1999; Aronu et al., 2011). It is generally recognized that this separation method offers high recoveries due to the high solvent selectivity and high methane concentration in the purified gas, but it still suffers the penalty of the energy consumed for the solvent regeneration. In order to keep the benefit of high solvent selectivity and to reduce the drawback of the substantial energy consumption, the hybrid membraneabsorption system was proposed by Research Institute of Innovative Technology for Earth and Taiyo Nippon Sanso Corporation (Tomioka et al., 2013). The hybrid process is based on the concept of bulk flow liquid membrane (Teramoto et al., 2003) reported in Fig. 6.17.

Figure 6.17 Representation of bulk flow liquid membrane.

In this arrangement, the solvent is supplied together with the feed gas to the high-pressure side of the ed liquid membrane. The solvent reacts with the carbon dioxide and permeates to the low-pressure side of the membrane. Due to the pressure difference the carbon dioxide is released, and the regenerated solvent can be recycled to the feed side after compression. This system was tested by Teramoto et al. (2001) for CO2/CH4 separation using an ultrafiltration polyethersulfone membrane and water as solvent. They observed a carbon dioxide permeability 20 times higher compared to the case without solvent permeation. The potential of this technology in regenerating the solvent with a minimum energy request represents an ideal match to overcome the absorberstripper energy drawback. Coupling the two-unit operations, it is possible to obtain the hybrid flowsheet of Fig. 6.18. The biogas is fed at the bottom of the absorber where the carbon dioxide is transferred to the liquid phase by reaction with the solvent and the purified methane stream is recovered from the top. The solvent and the carbon dioxide “rich solvent” are sent to the membrane unit where they are regenerated and sent back to the absorber. The stream rich in carbon dioxide is further purified and stored.

Figure 6.18 Hybrid membrane-absorption flowsheet.

The hybrid configuration was tested by Tomioka et al. (2013) using diethanolamine as a solvent for a 10 N m³/h unit obtaining a concentration of methane of 98 vol.% with a recovery that approached 98%. The purity of the carbon dioxide recovered was reported to be in the range 96–99 vol.%. Also, in this case, studies of hybrid membrane-absorption processes are limited but the results discussed highlighted how membrane-aided processes are becoming increasingly popular for the reduction of the energy request maintaining at the same time high purity targets. In a pilot-scale operation set up by the company Bioscan A/S at a medium-sized biogas plant in Denmark, CMSs were tested in a two-stage configuration. The membranes were supplied from CML Ltd, a company in Israel which is now closed, and had different operating characteristics: One had large permeability but low selectivity, the other had the opposite properties. The lay-out is shown in Fig. 6.19. The biogas is compressed to 0.7 bar (g) in a Roots blower to achieve the level of pressure required for a pressure swing adsorption process with zeolites removing residual water from the gas. Subsequently, the gas is compressed up to around 7 bar (g) and ed through an active carbon bed to remove residual hydrogen sulfide. The feed gas is then composed of methane, carbon dioxide, and less than 2% N2, which originates from prior hydrogen sulfide removal in a biological sulfide oxidizer unit. In the two-stage membrane configuration arranged as a cascade, the separation takes place in which the first module has a fairly large permeability but relatively low selectivity, hence giving a fairly high stage-cut, that is, high permeate flow with about 60% CO2, 38% CH4, and most of the nitrogen. In the next stage, the permeability is lower for CO2 and the selectivity high, giving a high concentration of CH4 of about 96% with the rest comprising CO2 and N2. The permeate from this stage is recycled to the feed. In a simulation, a cryogenic unit was inserted receiving the CO2-rich permeate stream from the first stage and cooling it down to about –37°C. The methane solubility in the liquid CO2 made in the cryogenic separator is so low that practically all the methane is flashed off from the unit together with nitrogen. The CO2 from the cryogenic unit is

claimed to be of a purity of more than 99.5% by the constructing company Union Engineering A/S, given that the feed gas complied with a requirement of low water and H2S content. The process is described in more detail in Norddahl and du Preez (2007).

Figure 6.19 Hybrid upgrading process for biogas with a two-stage CMS membrane unit coupled to a cryogenic CO2/CH4 separator (Norddahl and du Preez, 2007).

A study performed by Scholz et al. (2013) compared different upgrading technologies with simulations performed on hybrid systems comprising absorption processes, cryogenic upgrading, and pure membrane processes. The conclusion was that: “Membrane hybrid processes in which gas permeation technology is combined with established gas separation techniques such as pressurized water scrubbing, amine absorption, and cryogenic separation are attractive for biogas upgrading. These hybrid processes were primarily investigated in of upgrading costs using Aspen Plus. Investment costs are ed for by Guthrie’s method for equipment cost estimation. Comparing the hybrid processes to the established separation techniques clearly shows that established upgrading processes would benefit from a combination with membrane technology.” This is further accentuated with a graphical representation of the cost comparisons as shown in Fig. 6.20

Figure 6.20 Comparison of the different hybrid processes in of the specific upgrading costs.

Here, the upgrading costs are presented with respect to the product flow rate. The data are shown for a feed flow rate of 1000 m³ (STP)/h. The filled symbols refer to published data for the amine and the pressurized water-scrubbing process (Weidner et al., 2008).

6.6.3 Microbial conversion of CO2 to CH4 on a membrane diff

In an unconventional use of microfiltration HF membranes, anaerobic hydrogenotrophic archaea attached to the outer surface of an HF are met with a stream of biogas mixed with hydrogen diffusing through the membrane material in a configuration as shown in Fig. 6.21

Figure 6.21 Hollow-fiber microbial reactor for conversion of CO2 to CH4 in a microbial film attached to the surface of the HFs.

The basic design for a full-scale application is shown in Fig. 6.22

Figure 6.22 Sketch of a basic design for a full-scale application of microbial conversion of biogas to methane by adding hydrogen.

This configuration is described in the PhD thesis Integration of Biomass and Wind Power for Biogas Enhancement and Upgrading via Hydrogen Assisted Anaerobic Digestion by Tirunehe (2017). In the thesis, Tirunehe argues that a configuration where the hollow-fiber membrane bioreactors are placed outside the biogas digestion (ex situ—configuration) has the best performance with respect to investment, cost of operation, and overall performance measured as conversion efficiency (amount of CO2 converted to methane per area of HF membranes and consumption of hydrogen for the conversion). In a small pilot operation, it was shown that a conversion of over 99% of the carbon dioxide to methane was possible with a consumption of equimolar amounts of hydrogen according to the reaction equation:

(6.16)

without recycling of the mixed gases.

6.7 Summary and outlook

In a modern process synthesis approach, it is desirable that sustainability criteria are applied at the initial stages of any design development, as well as control of how these criteria are fulfilled and their impact on process competitiveness. In the case of upgrading biogas to NG quality, a natural choice of sustainability criteria is whether the methods used lead to a product complying with regulations for NG in case the product, Biomethane, is meant either for injection in the common gas grid, to be used as motor fuel, or as a raw material for production of a commodity chemical, such as ammonia. The latter is a likely possibility if NG is phased out for the reason of avoiding use of fossil materials. Another important parameter is the cost of the product. The cost cannot differ too much from products made from traditional competitive methods, as it impacts with socioeconomic issues—another sustainability requirement. It is also desirable that the use of the product complies with a sustainability requirement, that is, considering whether biomethane is used as a rather low-cost commodity as fuel, or whether it should be used as a raw material for production of higher value commodities like ammonia or further used in general chemical products. This would reduce the need for using fossil fuels as raw materials, anticipating the need for new raw materials, when fossil raw materials are phased out. Methane (presently as NG, but replaceable with upgraded biogas) will play an increasing role in the production of short-chained olefins like ethylene as important basic products via the intermediate of methanol from synthesis gas (Amghizar et al., 2017). Whereas the indirect synthesis via methanol has already reached technical maturity, numerous questions still need to be addressed regarding selective olefin production from methane (Wang et al., 2010). Another area of interest is for the direct conversion of methane into aromatic compounds or functionalized products (Morejudo et al., 2016). Methane in this context requires compliance to the same standards as for grid injection. The sustainable production from upgraded biogas of polymers and solvents thus appears a possible future scenario. This has the potential to turn membrane

production into a more sustainable process also. As the use of biomass, especially biomass in the form of residual products from agriculture and food production, as feed for the biogas process is increasing so is the possibility of using methane to replace NG. The longstanding question is whether sustainability in this context complies with economics. In the earlier sections of this chapter, a number of hybrid membrane processes were described. In the processes described by Scholz et al. (2013), which is a model study based on empirical data with gas flows in the range of 150– 2000 N m³/h, data prices in the range of 0.14–0.075 EUR/N m³ were referenced and in the test set up by Bioscan (Norddahl and du Preez, 2007), which was based on results from a pilot plant in the biogas flow range 30 N m³/h, but extended to an economical model calculation with empirical data to 500 N m³/h, a price of around 0.08 EUR/N m³ for upgraded biogas in a hybrid membrane/cryogenic plant was calculated. In the statistical yearbook for EU 2019 (E-Yearbook, 2019) an average gas price for EU countries is stated as 0.671 EUR/N m³ (18.66 EUR/GJ and a lower heating value of 10 kWh/N m³ methane). The conclusion is straightforward: With a production price of hybrid upgraded biogas in the range of 0.14–0.075 EUR/N m³ biogas (Scholz et al., 2013) and an average price of NG of 0.671 EUR/N m³ with tax in the EU (E-Yearbook, 2019), the price of upgraded biogas is absolutely competitive with NG. Moreover, by using a hybrid process like the membrane/cryogenic one, where the CO2 separated from the biogas can be used and sold as an industry commodity, it is possible to increase the sustainability of the overall process.

6.8 Future developments

In the Paris Agreement on future climate change mitigation, the declared goal is: “Holding the increase in the global average temperature to well below 2°C above preindustrial levels and to pursue efforts to limit the temperature increase to 1.5°C above preindustrial levels, recognizing that this would significantly reduce the risks and impacts of climate change.” In a strict sense, this implies a gradual change from the use of fossil raw materials toward more sustainable raw materials for the production of both fuels and commodities. In this scenario, biogas and its upgrading through membrane technology has the potential to be the frontrunner technology. In particular, biogas enriched in methanol has an obvious application in the fuel sector, but even more interesting is its application for commodity chemicals production. This has an impact also in the membrane production processes. For the use of membranes in the future for gas separation, this implies the use of renewable organic raw materials like cellulose and cellulose derivatives and polymers made with sugar as the base or inorganic materials. This is definitely feasible, albeit, probably at an elevated cost compared to the present time. However, already CMSs are made from cellulose that uses wood as the raw material, and the technology for using sugar to produce polymers is well established, that is, by using genetically engineered microorganisms for production of 1,4-butanediol, which is a precursor for production of many polymers. Sugar can also be used as feed for microorganisms producing both succinic and fumaric acids, that are known precursors for many different polymers including polyethylene and polyamides and imides (Rehm, 2009). Another promising polymer made from plant-based material is poly-(ethylene 2,5-furandicarboxylate), a glassy polymer with prospects as a gas separation membrane. A review of biobased polymers is presented by Nakajima et al. (2017). Further to this, in the earlier sections it was described how the upgraded methane itself can be used as a precursor for almost any polymer given a sufficient number of intermediate steps are used.

Generally, it is evident that the technology for producing sustainable gas separation membranes from renewable materials is present and available. Thus, in a future governed by nonfossil resources, it will be possible to upgrade biogas to NG quality with membranes.

References

Ahmad et al., 2018 Ahmad MZ, et al. Investigation of a new co-polyimide, 6FDA-bisP and its ZIF-8 mixed matrix membranes for CO2/CH4 separation. Sep Purif Technol. 2018;207:523–534. Alentiev and Yampolskii, 2000 Alentiev AY, Yampolskii YP. Free volume model and tradeoff relations of gas permeability and selectivity in glassy polymers. J Membr Sci. 2000;165(2):201–216. Álvarez et al., 2020 Álvarez C, et al. Gas separation properties of aromatic polyimides with bulky groups Comparison of experimental and simulated results. J Membr Sci. 2020;117959. Amghizar et al., 2017 Amghizar I, et al. New trends in olefin production. Engineering. 2017;3(2):171–178. Angelidaki et al., 2018 Angelidaki I, et al. Biogas upgrading and utilization: current status and perspectives. Biotechnol Adv. 2018;36(2):452–466. Aronu et al., 2011 Aronu UE, Hoff KA, Svendsen HF. CO2 capture solvent selection by combined absorption–desorption analysis. Chem Eng Res Des. 2011;89(8):1197–1203. Baccioli et al., 2018 Baccioli A, et al. Small scale bio-LNG plant: comparison of different biogas upgrading techniques. Appl Energy. 2018;217:328–335. Baker, 2002 Baker RW. Future directions of membrane gas separation technology. Ind Eng Chem Res. 2002;41(6):1393–1411. Baker, 2004 Baker RW. Membrane Technology and Applications John Wiley & Sons, Ltd 2004;96–103. Baker and Lokhandwala, 2008 Baker RW, Lokhandwala K. Natural gas processing with membranes: an overview. Ind Eng Chem Res. 2008;47(7):2109–

2121. Baker and Low, 2014 Baker RW, Low BT. Gas separation membrane materials: a perspective. Macromolecules. 2014;47(20):6999–7013. Bastani et al., 2013 Bastani D, Esmaeili N, Asadollahi M. Polymeric mixed matrix membranes containing zeolites as a filler for gas separation applications: a review. J Ind Eng Chem. 2013;19(2):375–393. Bernardo and Drioli, 2010 Bernardo P, Drioli E. Membrane gas separation progresses for process intensification strategy in the petrochemical industry. Pet Chem. 2010;50(4):271–282. Bernardo et al., 2009 Bernardo P, Drioli E, Golemme G. Membrane gas separation: a review/state of the art. Ind Eng Chem Res. 2009;48(10):4638– 4663. Bos et al., 1998 Bos A, et al. Plasticization-resistant glassy polyimide membranes for CO2/CO4 separations. Sep Purif Technol. 1998;14(1-3):27–39. Brunetti et al., 2010 Brunetti A, et al. Membrane technologies for CO2 separation. J Membr Sci. 2010;359(1-2):115–125. Burdyny and Struchtrup, 2010 Burdyny T, Struchtrup H. Hybrid membrane/cryogenic separation of oxygen from air for use in the oxy-fuel process. Energy. 2010;35(5):1884–1897. Chen et al., 2012 Chen X, Rodrigue D, Kaliaguine S. Diamino-organosilicone APTMDS: a new cross-linking agent for polyimides membranes. Sep Purif Technol. 2012;86:221–233. Chen et al., 2015 Chen XY, et al. Membrane gas separation technologies for biogas upgrading. RSC Adv. 2015;5(31):24399–24448. Chen et al., 2017 Chen, X.Y., Tien-Binh, N., Kaliaguine, S., Rodrigue, S., 2017. Polyimide membranes for gas separation: synthesis, processing and properties. In: Murphy, N.P.C. (Ed.), Polyimides Synthesis, Applications and Research, New York. Commission, 2011 Commission EE. Energy roap 2050. COM. 2011;885:15

(2011). Delgado et al., 2009 Delgado JA, et al. Simulation of CO2 absorption into aqueous DEA using a hollow fiber membrane or: evaluation of or performance. Chem Eng J. 2009;152(2-3):396–405. Doghieri et al., 2006 Doghieri F, et al. Solubility of gases and vapors in glassy polymers modelled through non-equilibrium PHSC theory. Fluid Phase Equilibria. 2006;241(1-2):300–307. Dortmundt and Doshi, 1999 Dortmundt D, Doshi K. Recent developments in CO2 removal membrane technology. UOP LLC 1999;1. Drioli, 2002 Drioli, E., 2002. An international report on Membranes Science and Technology, perspectives and needs. In: Prepared and discussed at the International Conference on Membranes (ICOM 2002) held in Toulouse on July. Drioli and Giorno, 2009 Drioli E, Giorno L. Membrane Operations: Innovative Separations and Transformations John Wiley & Sons 2009. Errico, 2017 Errico M. Process synthesis and intensification of hybrid separations. Process Synthesis and Process Intensification De Gruyter 2017;182– 212. Esposito et al., 2015 Esposito E, et al. Pebax®/PAN hollow fiber membranes for CO2/CH4 separation. Chem Eng Process.—Process Intensif. 2015;94:53–61. Estahbanati et al., 2017 Estahbanati EG, Omidkhah M, Amooghin AE. Preparation and characterization of novel ionic liquid/Pebax membranes for efficient CO2/light gases separation. J Ind Eng Chem. 2017;51:77–89. Farashi et al., 2019 Farashi Z, et al. Improving CO2/CH4 separation efficiency of Pebax-1657 membrane by adding Al2O3 nanoparticles in its matrix. J Nat Gas Sci Eng. 2019;72:103019. Felder and Huvard, 1980 Felder R, Huvard G. 17 Permeation, diffusion, and sorption of gases and vapors. Methods in Experimental Physics Elsevier 1980;315–377. Freeman, 1999 Freeman BD. Basis of permeability/selectivity tradeoff relations

in polymeric gas separation membranes. Macromolecules. 1999;32(2):375–380. Freeman et al., 2006 Freeman B, Yampolskii Y, Pinnau I. Materials Science of Membranes for Gas and Vapor Separation John Wiley & Sons 2006. Graham, 1833 Graham T. On the law of the diffusion of gases (continued). Phil Mag. 1833;2:354. Guerrero et al., 2015 Guerrero L, et al. Advances in the biological removal of sulphides from aqueous phase in anaerobic processes: a review. Environ Rev. 2015;24(1):84–100. Hägg and He, 2011 Hägg, M.-B., He X., 2011. Chapter 15. Carbon Molecular Sieve Membranes for Gas Separation. 162–191. Haider et al., 2018 Haider S, et al. Pilot(-)scale production of carbon hollow fiber membranes from regenerated cellulose precursor-part II: carbonization procedure. Membrane (Basel). 2018;8. Hamid and Jeong, 2018 Hamid MRA, Jeong H-K. Recent advances on mixedmatrix membranes for gas separation: opportunities and engineering challenges. Korean J Chem Eng. 2018;35(8):1577–1600. Harasimowicz et al., 2007 Harasimowicz M, et al. Application of polyimide membranes for biogas purification and enrichment. J Hazard Mater. 2007;144(3):698–702. He, 2018 He X. A review of material development in the field of carbon capture and the application of membrane-based processes in power plants and energyintensive industries. Energy Sustainability Soc. 2018;8(1):34. He et al., 2018 He X, et al. Carbon molecular sieve membranes for biogas upgrading: techno-economic feasibility analysis. J Clean Prod. 2018;194:584– 593. Heile et al., 2014 Heile S, et al. Establishing the suitability of symmetric ultrathin wall polydimethylsiloxane hollow-fibre membrane ors for enhanced CO2 separation during biogas upgrading. J Membr Sci. 2014;452:37– 45.

Henis and Tripodi, 1980 Henis JM, Tripodi MK. A novel approach to gas separations using composite hollow fiber membranes. Sep Sci Technol. 1980;15(4):1059–1068. Hjuler Klaus, 2017a Hjuler Klaus, A.N., 2017a. Review on Biogas Upgrading. Project Report, D.G.C. A/S, Editor. Hjuler Klaus, 2017b Hjuler Klaus, A.N., 2017b. Review on Biogas Upgrading, D.G.T. Centre, Editor. Hoehn, 1985 Hoehn H. Aromatic Polyamide Membranes ACS Publications 1985. Ismail et al., 2015 Ismail AF, Khulbe KC, Matsuura T. Gas Separation Membranes Switzerland: Springer; 2015. Jeon and Lee, 2015 Jeon Y-W, Lee D-H. Gas membranes for CO2/CH4 (biogas) separation: a review. Environ Eng Sci. 2015;32(2):71–85. Kapoor et al., 2019 Kapoor R, et al. Evaluation of biogas upgrading technologies and future perspectives: a review. Environ Sci Pollut Res Int. 2019;26(12):11631–11661. Koros and Paul, 1981 Koros W, Paul D. Observations concerning the temperature dependence of the Langmuir sorption capacity of glassy polymers. J Polym Sci.: Polym Phys Ed. 1981;19(10):1655–1656. Koros and Chern, 1987 Koros, W.J., Chern R.T., 1987. Separation of gaseous mixtures using polymer membranes. Handbook of Separation Process Technology, pp. 862–953. Koros and Fleming, 1993 Koros WJ, Fleming G. Membrane-based gas separation. J Membr Sci. 1993;83(1):1–80. Koros et al., 1977 Koros WJ, Chan A, Paul DR. Sorption and transport of various gases in polycarbonate. J Membr Sci. 1977;2:165–190. Lagorsse, 2004 Lagorsse S. Carbon molecular sieve membranes sorption, kinetic and structural characterization. J Membr Sci. 2004;241(2):275–287.

Liao et al., 2018 Liao Z, et al. Optimal design of hybrid cryogenic flash and membrane system. Chem Eng Sci. 2018;179:13–31. Lin and Freeman, 2005 Lin H, Freeman BD. Materials selection guidelines for membranes that remove CO2 from gas mixtures. J Mol Struct. 2005;739(13):57–74. Liu et al., 2013 Liu L, et al. Influence of membrane skin morphology on CO2/N2 separation at sub-ambient temperatures. J Membr Sci. 2013;446:433– 439. Madeddu et al., 2018 Madeddu C, Errico M, Baratti R. Process analysis for the carbon dioxide chemical absorption–regeneration system. Appl Energy. 2018;215:532–542. Mahmoudi et al., 2015 Mahmoudi A, Asghari M, Zargar V. CO2/CH4 separation through a novel commercializable three-phase PEBA/PEG/NaX nanocomposite membrane. J Ind Eng Chem. 2015;23:238–242. Makaruk et al., 2010 Makaruk A, Miltner M, Harasek M. Membrane biogas upgrading processes for the production of natural gas substitute. Sep Purif Technol. 2010;74(1):83–92. Makaruk et al., 2013 Makaruk A, Miltner M, Harasek M. Biogas desulfurization and biogas upgrading using a hybrid membrane system–modeling study. Water Sci Technol. 2013;67(2):326–332. Maqsood et al., 2014 Maqsood K, et al. Cryogenic carbon dioxide separation from natural gas: a review based on conventional and novel emerging technologies. Rev Chem Eng. 2014;30(5):453–477. Matteucci et al., 2006 Matteucci S, et al. Transport of gases and vapors in glassy and rubbery polymers. Mater Sci Membr Gas Vap Sep. 2006;1:1–2. Mazzella et al., 2016 Mazzella A, Errico M, Spiga D. CO2 uptake capacity of coal fly ash: influence of pressure and temperature on direct gas-solid carbonation. J Environ Chem Eng. 2016;4(4):4120–4128. Minelli and Sarti, 2013 Minelli M, Sarti GC. Permeability and diffusivity of CO2 in glassy polymers with and without plasticization. J Membr Sci.

2013;435:176–185. Moore et al., 2004 Moore TT, et al. Hybrid membrane materials comprising organic polymers with rigid dispersed phases. AIChE J. 2004;50(2):311–321. Morejudo et al., 2016 Morejudo S, et al. Direct conversion of methane to aromatics in a catalytic co-ionic membrane reactor. Science. 2016;353(6299):563–566. Mourgues and Sanchez, 2005 Mourgues A, Sanchez J. Theoretical analysis of concentration polarization in membrane modules for gas separation with feed inside the hollow-fibers. J Membr Sci. 2005;252(1-2):133–144. Nagai et al., 2001 Nagai K, et al. Poly [1-(trimethylsilyl)-1-propyne] and related polymers: synthesis, properties and functions. Prog Polym Sci. 2001;26(5):721– 798. Nakajima et al., 2017 Nakajima H, Dijkstra P, Loos K. The recent developments in biobased polymers toward general and engineering applications: polymers that are upgraded from biodegradable polymers, analogous to petroleum-derived polymers, and newly developed. Polymers. 2017;9(10):523. Nakhjiri et al., 2018 Nakhjiri AT, et al. Experimental investigation and mathematical modeling of CO2 sequestration from CO2/CH4 gaseous mixture using MEA and TEA aqueous absorbents through polypropylene hollow fiber membrane or. J Membr Sci. 2018;565:1–13. Norddahl and du Preez, 2007 Norddahl, B., du Preez J., 2007. A membrane based process for the upgrading of biogas to substituted natural gas (SNG) and recovery of carbondioxide for industrial use. In: Proceedings of European Congress of Chemical Engineering (ECCE-6). Ockwig and Nenoff, 2007 Ockwig NW, Nenoff TM. Membranes for hydrogen separation. Chem Rev. 2007;107(10):4078–4110. Ozturk and Demirciyeva, 2013 Ozturk B, Demirciyeva F. Comparison of biogas upgrading performances of different mixed matrix membranes. Chem Eng J. 2013;222:209–217. Park et al., 2007 Park HB, et al. Polymers with cavities tuned for fast selective

transport of small molecules and ions. Science. 2007;318(5848):254–258. Park et al., 2017 Park A, et al. Biogas upgrading using membrane or process: pressure-cascaded stripping configuration. Sep Purif Technol. 2017;183:358–365. Pechar et al., 2006 Pechar T, et al. Fabrication and characterization of polyimide–zeolite L mixed matrix membranes for gas separations. J Membr Sci. 2006;277(1-2):195–202. Peppers et al., 2019 Peppers J, et al. Performance analysis of membrane separation for upgrading biogas to biomethane at small scale production sites. Biomass Bioenergy. 2019;128. Pixton and Paul, 1994 Pixton, M., Paul D., 1994. Polymeric gas separation membranes. Paul, DR, pp. 83–153. Prangsgaard Andersen, 2016 Prangsgaard Andersen, T., 2016. Hollow Fiber Carbon Membranes for Separation of Hydrogen and Carbon Dioxide (PhD thesis) in University of Southern Denmark. Raboni et al., 2015 Raboni M, Viotti P, Capodaglio AG. A comprehensive analysis of the current and future role of biofuels for transport in the European Union (EU) Ambiente e Agua—An Interdisciplinary. J Appl Sci. 2015;10. Rehm, 2009 Rehm BH. Microbial Production of Biopolymers and Polymer Precursors: Applications and Perspectives Horizon Scientific Press 2009. Robeson, 2008 Robeson LM. The upper bound revisited. J Membr Sci. 2008;320(1-2):390–400. Sanders et al., 2013 Sanders DF, et al. Energy-efficient polymeric gas separation membranes for a sustainable future: a review. Polymer. 2013;54(18):4729–4761. Saufi and Ismail, 2004 Saufi S, Ismail A. Fabrication of carbon membranes for gas separation––a review. Carbon. 2004;42(2):241–259. Schell, 1983 Schell W. Membrane use/technology growing. Hydrocarbon Process (USA). 1983;62.

Scholz et al., 2013 Scholz M, et al. Techno-economic analysis of hybrid processes for biogas upgrading. Ind Eng Chem Res. 2013;52(47):16929–16938. Soffer et al., 1999 Soffer, A., et al., 1999. Process for the production of hollow carbon fiber membranes. Google Patents. Song et al., 2017 Song C, et al. Reducing the energy consumption of membranecryogenic hybrid CO2 capture by process optimization. Energy. 2017;124:29– 39. Song et al., 2018a Song C, et al. Efficient biogas upgrading by a novel membrane-cryogenic hybrid process: experiment and simulation study. J Membr Sci. 2018a;565:194–202. Song et al., 2018b Song C, et al. Parametric study of a novel cryogenicmembrane hybrid system for efficient CO2 separation. Int J Greenh Gas Control. 2018b;72:74–81. Spillman, 1995 Spillman R. Economics of gas separation membrane processes. Membrane Science and Technology Elsevier 1995;589–667. Sridhar et al., 2007 Sridhar S, et al. Gas permeation properties of polyamide membrane prepared by interfacial polymerization. J Mater Sci. 2007;42(22):9392–9401. Stern et al., 1987 Stern S, Shah V, Hardy B. Structure-permeability relationships in silicone polymers. J Polym Sci Part B: Polym Phys. 1987;25(6):1263–1298. Sun et al., 2015 Sun Q, et al. Selection of appropriate biogas upgrading technology—a review of biogas cleaning, upgrading and utilisation. Renew Sustain Energy Rev. 2015;51:521–532. Teramoto et al., 2001 Teramoto M, et al. Gas separation by liquid membrane accompanied by permeation of membrane liquid through membrane physical transport. Sep Purif Technol. 2001;24(1-2):101–112. Teramoto et al., 2003 Teramoto M, et al. Separation and enrichment of carbon dioxide by capillary membrane module with permeation of carrier solution. Sep Purif Technol. 2003;30(3):215–227.

Tirunehe, 2017 Tirunehe, G., 2017. SYMBIO: Integration of Biomass and Wind Power for Biogas Enhancement and Upgrading via Hydrogen Assisted Anaerobic Digestion (PhD thesis), University of Southern Denmark. Tomioka et al., 2013 Tomioka T, Sakai T, Mano H. Carbon dioxide separation technology from biogas by “Membrane/Absorption Hybrid Method. Energy Procedia. 2013;37:1209–1217. Ullah Khan et al., 2017 Ullah Khan I, et al. Biogas as a renewable energy fuel— a review of biogas upgrading, utilisation and storage. Energy Convers Manag. 2017;150:277–294. Vega et al., 2019 Vega J, et al. Modification of polyetherimide membranes with ZIFs fillers for CO2 separation. Sep Purif Technol. 2019;212:474–482. Vrbová and Ciahotný, 2017 Vrbová V, Ciahotný K. Upgrading biogas to biomethane using membrane separation. Energy Fuels. 2017;31(9):9393–9401. Vu et al., 2003 Vu DQ, Koros WJ, Miller SJ. Mixed matrix membranes using carbon molecular sieves: I Preparation and experimental results. J Membr Sci. 2003;211(2):311–334. Wahab and Sunarti, 2015 Wahab MA, Sunarti A. Development of PEBAX based membrane for gas separation: A. Int J. 2015;2(2):79. Wang et al., 2010 Wang Y, Wu R, Zhao Y. Effect of ZrO2 promoter on structure and catalytic activity of the Ni/SiO2 catalyst for CO methanation in hydrogenrich gases. Catal Today. 2010;158(3-4):470–474. WBA, 2018 WBA, 2018. W.B.A., WBA Global Bioenergy Statistics 2018 Summary Report. Weidner et al., 2008 Weidner, E., et al., 2008. Technologien und Kosten der Biogasaufbereitung und Einspeisung in das Erdgasnetz. Ergebnisse der Markterhebung 2007–2008. Institut Umwelt-, Sicherheits-, Energietechnik. Wijmans and Baker, 1995 Wijmans JG, Baker RW. The solution-diffusion model: a review. J Membr Sci. 1995;107(1-2):1–21. Yampolskii, 2007 Yampolskii YP. Methods for investigation of the free volume

in polymers. Russian Chem Rev. 2007;76(1):59. Yampolskii, 2012 Yampolskii Y. Polymeric gas separation membranes. Macromolecules. 2012;45(8):3298–3311. Yan et al., 2014 Yan S, et al. Biogas upgrading by CO2 removal with a highly selective natural amino acid salt in gas–liquid membrane or. Chem Eng Process.: Process Intensif. 2014;85:125–135. E-Yearbook, 2019 E-Yearbook, 2019. Retrieved on 26.11.2019. . Žák et al., 2018 Žák M, et al. Single-step purification of raw biogas to biomethane quality by hollow fiber membranes without any pretreatment—an innovation in biogas upgrading. Sep Purif Technol. 2018;203:36–40. Zhang and Wang, 2013 Zhang Y, Wang R. Gas–liquid membrane ors for acid gas removal: recent advances and future challenges. Curr Opin Chem Eng. 2013;2(2):255–262. Zhang et al., 2018 Zhang Z, et al. Zeolites nanocomposite membrane applications in CO2 capture. Handbook of Nanomaterials for Industrial Applications Elsevier 2018;916–921.

Chapter 7

Cryogenic techniques: an innovative approach for biogas upgrading

Francisco Manuel Baena-Moreno, Luz M. Gallego, Fernando Vega and Benito Navarrete, Chemical and Environmental Engineering Department, Technical School of Engineering, University of Seville, Sevilla, Spain

Abstract

In this chapter cryogenic techniques as an alternative to biogas upgrading are analyzed in depth. The growth of biogas production needs an evolution toward new biogas upgrading techniques which allow obtaining liquefied biogas (LBG). In this sense, cryogenic techniques are a promising candidate for biogas upgrading in places where an excess of cool streams is available. Cryogenic techniques provide good separation of all the components present in biogas, as well as high purity of the final biomethane and CO2, which can be sold to balance the overall costs. The main problem with cryogenic technologies is the energy consumption. Therefore research efforts are currently aiming to reduce this aspect. This work is provided as a guide to those involved in this path regarding cryogenic techniques. To this end, the traditional cryogenic techniques (cryogenic distillation and cryogenic packed-bed technology) are explained and the most relevant works to date are reviewed. Moreover, the combination of cryogenic techniques with other biogas upgrading technologies is presented as innovative integrations for maximizing the benefits of each technology. Thus cryogenic combinations with absorption, adsorption, membranes, and hydrate technologies are analyzed. Among these combinations, cryogenic-membrane technology is highlighted as it allows to reduce the energy consumption of the overall system and to obtain high-purity final streams. Future works in this area should lead to scaling up of the different configurations herein described and achieve more realistic data which could help in the development of cryogenic technologies at industrial scales.

Keywords

Biogas upgrading; cryogenic techniques; hybrid cryogenic systems; innovative cryogenic approaches

Chapter outline

Outline

7.1 Introduction 159

7.2 Cryogenic biogas upgrading 160

7.2.1 Cryogenic distillation 162 7.2.2 Cryogenic packed-bed technology 163

7.3 Cryogenic hybrid systems 166

7.3.1 Cryogenic-absorption combination process 167 7.3.2 Cryogenic-adsorption synergized process 168 7.3.3 Potential combination of cryogenic and membrane processes 170 7.3.4 Cryogenic-hydrate processes 170

7.4 Cryogenic-membrane processes 171

7.5 Full-scale experiences and technoeconomic studies 176

7.6 Comparison of documented technologies 177

7.7 Conclusions and future perspectives 179

Appendix I: Conversion factor for unit transformations 180

Appendix II: State forms for CO2 and CH4 as a function of temperature and pressure 180

Acknowledgments 181

References 181

7.1 Introduction

The search for novel solutions to tackle the fossil fuel energy dependence is an ongoing topic for the research community. In this sense, renewable energy obtained from biogas is a promising option. Biogas consists basically of a mixture of CH4 (60%) and CO2 (40%), which comes from the anaerobic digestion of waste (Angelidaki et al., 2018). To obtain a 100% green source of energy, CO2 needs to be removed before biogas utilization (Baena-Moreno et al., 2019a). The product stream (without CO2) is known as biomethane and the technique to remove CO2 is called biogas upgrading. The techniques for biogas upgrading can be classified into chemical absorption, physical absorption, membrane technologies, and cryogenic upgrading (Sun et al., 2015). During the last few years, the many authors have tried to find a utilization path for the captured CO2 and hence make the process affordable or more competitive from an economic perspective (Baena-Moreno et al., 2018a). Following this integration, two benefits can be obtained. First, a biomethane stream is obtained as a possible fuel or green energy source. Moreover, added-value materials are produced from contaminants such as CO2. Indeed, several works have been conducted during the last few years to fulfill this integration (Baena-Moreno et al., 2019b,c,d; Horschig et al., 2019; Garcia-Herrero et al., 2016). These works have been mainly led to the production of carbamates or carbonates (BaenaMoreno et al., 2018b, 2019e), or the production of formic acid and methanol as commodities (Baena-Moreno et al., 2020a,b). In as much as CO2 could be also used directly (without chemical transformation), integrations between biogas upgrading and CO2 utilization which allow the direct utilization of CO2 must be explored (Baena-Moreno et al., 2020c; Zhang et al., 2020). In this sense, cryogenic technology offers a promising alternative due to the state in which CO2 is obtained (Pellegrini et al., 2018). Moreover, a high-purity CH4 stream is obtained after the cryogenic operation. The CH4 losses caused in cryogenic technology are controllably burnt in order to maintain the GHG balance. Even though many cryogenic alternatives are available for CO2 separation from flue gas, only a few studies propose this application to upgrade biogas. Therefore this synergy offers new research

opportunities. Thus guidelines for those exploring this technology is needed. This chapter aims to provide reference guidelines for the research teams immersed in this topic. For a proper comprehension of cryogenic techniques employment for biogas upgrading and CO2 separation, this chapter is organized as follows. First, cryogenic biogas upgrading is explained from the basics to the most relevant parts of the technology. Then, cryogenic hybrid systems, which are aimed at obtaining the best characteristics of each technology, are explained. As cryogenic-membrane processes are the most promising option among these combinations, the final section is dedicated to this configuration.

7.2 Cryogenic biogas upgrading

Cryogenic techniques can be a good option to produce high-purity products or higher fuel standards from a flue gas or a biogas stream, when different compounds such as SH2, CO2, CH4, O2, and N2, show variations in their condensation temperatures (Yousef et al., 2018). Compared with other biogas upgrading techniques such as water scrubbing, physical–chemical absorption, or membrane techniques (Sun et al., 2015), cryogenic approaches have the benefits listed below (Bauer et al., 2013):

1. Indirect between gas and chemicals; 2. Production of high-purity CO2 as a secondary product; 3. Good separation of all the components present in biogas; 4. Production of liquefied biogas (LBG); 5. Possibility to remove nitrogen from the gas stream.

Currently, the main challenge to cryogenic technology is the high energy requirement associated with these purification processes. This fact prevents its expansion into the commercial field (Bauer et al., 2013). Fig. 7.1 shows as an example the GPP system of gas treatment services (GtS), in which the efficiency of the CH4 recovery is higher than 98% and the operational costs are extremely low since the system only requires electricity for cooling (Energy GF, 2019). In this system, biogas is first filtered and dried, followed by compression to adjust the pressure requirements. Then, the biogas is cleaned in two stages as shown in Fig. 7.1, with a defrost step to avoid frosting between the stages.

Figure 7.1 Scheme of the GPP system of gas treatment services for biogas upgrading. Adapted from Energy GF. Biogas upgrading. n.d. <www.gtsbv.com/over-gts/biogas-opwaardering> (accessed 16.11.19).

Technologies based on cryogenic separation processes are more energy-efficient to upgrade large volumes of gases with high-concentration CO2. Traditional methods such as membrane separation or water scrubbing use multiple stages and are not as effective for CO2 removal (Yousef et al., 2018). For this reason, one of the most promising alternatives is the combination of a cryogenic process with carbon capture and utilization (Yousef et al., 2018).

7.2.1 Cryogenic distillation

Cryogenic distillation can be used to remove trace contaminants from biogas produced in anaerobic digestion processes or landfills (N2 or O2 from the CH4), or to separate the CH4 from CO2 streams and subsequently liquefy it to produce bio-liquefied natural gas (bio-LNG), or LBG (Benjaminsson et al., 2013). This process harnesses the different boiling points that CO2 and CH4 have (−164°C for CH4 and −78°C for CO2 at 1 bar), so that distillation is feasible when the biogas has been cooled to these low temperatures. Fig. 7.2 shows the components and units used for low-temperature CH4/CO2 separation. Among these units, the compression stage, distillation column, and the final flash for adjusting the quality are highlighted. The minor impurities of raw biogas are removed in a preconditioning stage, such as H2O and N2 (Yousef et al., 2019).

Figure 7.2 Cryogenic distillation process scheme. Adapted from Yousef A.M., El-Maghlany W.M., Eldrainy Y.A., Attia A., 2019. Upgrading biogas to biomethane and liquid CO2: a novel cryogenic process. Fuel. 251, https://doi.org/10.1016/j.fuel.2019.03.127.

The separation process or CO2 liquefaction is carried out in four differentiated steps:

1. Compression at high pressures with intermediate water refrigeration to increase the pressure up to 4900 kPa and the biogas temperature around 74°C; 2. Deep cooling using a precooler heat exchanger and cooler working on a refrigeration cylcle to reduce the temperature of the gas stream to −40°C, keeping the pressure value; 3. Distillation and liquefaction using a distillation column, which generates from the top a gas stream at 95.5% of CH4, and from the bottom a valuable CO2 byproduct liquid stream at 99.3%; 4. Flashing the gas stream generated at the top of the column, and purifying to achieve 97.2% of CH4.

The principal obstacle hindering the implementation of this technology is the CO2 freezing out along the system, causing frosting and increasing energy consumption in the distillation column (Yousef et al., 2018; Zhang and Lior, 2006, 2008; Zhang et al., 2010; Ali et al., 2010). There have been many studies aimed at achieving an upgraded biogas with 97.12% CH4 purity and a byproduct with 99.92% CO2 purity which avoid these limitations. The distillation pressure, temperature, reflux ratio, and number of trays are varied through the implementation of a second distillation column (Yousef et al., 2017, 2018). Fig. 7.3 shows the low-temperature distillation model proposed by Yousef et al. (2018), in which can be seen many changes in comparison with the basic scheme (Fig. 7.2). The main novelty is the incorporation of a second distillation column

which allows for further adjustment of the biogas quality required.

Figure 7.3 Cryogenic distillation process scheme investigated to avoid frosting and lowering the energy consumption in the separation process. Adapted from Yousef A.M., Eldrainy Y.A., El-Maghlany W.M., Attia A. 2017. Biogas upgrading process via low-temperature CO2 liquefaction and separation. J. Nat. Gas Sci. Eng. 45, https://doi.org/10.1016/j.jngse.2017.07.001.

This configuration allows reducing the specific energy demand from 0.8– 1.5 kWh/Nm³ of cleaned gas (traditional cryogenic separation, Ali et al., 2010) to 0.5–0.6 kWh/Nm³ of cleaned gas. Table 7.1 presents a comparison between the aforementioned process and the traditional biogas upgrading approaches in of specific energy demand. These new findings could make cryogenic distillation applicable and competitive in comparison with traditional biogas upgrading approaches.

Table 7.1

Biogas upgrading techniques

Specific energy demand (kWh/Nm³ cleaned gas)

Cryogenic distillation (2 distillation columns)

0.5–0.6

Water scrubbing

0.45–0.92

Physical absorption

0.5–0.7

Chemical absorption

0.4–0.45

Pressure swing adsorption

0.3–1

Membrane technology

0.25–0.42

Typical cryogenic separation (one distillation column)

0.8–1.55

Adapted from Yousef A.M., El-Maghlany W.M., Eldrainy Y.A., Attia A., 2019. Upgrading biogas to biomethane and liquid CO2: a novel cryogenic process. Fuel. 251, https://doi.org/10.1016/j.fuel.2019.03.127.

7.2.2 Cryogenic packed-bed technology

Cryogenic packed bed (B) has been highly developed to capture CO2 from flue gases from the industry, but this technology is increasingly applied to a wide range of other gas separations such as biogas upgrading. This process is based on dynamic operations of adsorption stages where packed beds allow operation in different processes, such as cooling, capture, and recovery steps. In B the bed is cooled to temperatures lower than −100°C and pressures higher than 40 bars, where the contaminants in the biogas (CO2 and H2O) are condensed/sublimated, and other compounds (CH4 and N2) remain in the gas stream. For example, when feeding a mix flue gas of CH4 and CO2 to a previously refrigerated packed bed, the flue gas will cool down and the packing bed will heat up until the CO2 starts to desublimate at the packaging surface and move toward the outlet to the bed (Fig. 7.4) (Tuinier et al., 2010). Then, the bed is switched to a recovery stage.

Figure 7.4 CO2 ice formed at the packing surface during a capture cycle. Adapted from Tuinier M.J., van Sint Annaland M., Kramer G.J., Kuipers J.A.M. 2010. Cryogenic CO2 capture using dynamically operated packed beds. Chem. Eng. Sci. 65, https://doi.org/10.1016/j.ces.2009.01.055.

In contrast to the usual capture process, for biogas treatment CO2 capture is not required. Therefore N2 or air can be used in the recovery step, although when using air, CH4 and O2 might form mixtures which are within explosion limits. As the amount of CH4 present in the gas phase of the bed is limited, the amount of N2 required is also minimal. When the CH4 is removed from the bed, the recovery can be carried out using air, and the mixing of CH4 and air is avoided. As the amount of cold energy stored in the packing is restricted, the quantity of CO2 deposition is limited, hence avoiding plugging in the units. This is an advantage in comparison with the use of conventional cryogenic technologies concerning freezing out CO2 from gas streams. Furthermore, the continuous cooling given in the traditional technologies reduces the heat transfer rate toward the gas phase due to the increased amount of solid CO2 at the heat exchanging surfaces. Another advantage, together with the simple construction of the beds, is that maximum CO2 removal is guaranteed. This is because the gas stream leaving the bed is at the same temperature as the initial bed temperature (zero approach temperature). Several studies have been performed in order to improve the energy consumption and CH4 losses of this process by varying the temperature, pressure, and raw gas composition (Tuinier et al., 2011a; DiMaria et al., 2015; Ali et al., 2018). From the positive side, these works show that it is possible to achieve losses below 6% and therefore achieve a product purity of 94%. Moreover, it is likely that an acceptable and competitive energy consumption rate for biogas improvement can be obtained, as demonstrated in the study by Tuinier et al. (2011a). In this research, B technology was tested at a coal power plant, where an energy consumption rate of 1.8 MJ/kg CO2 was achieved. These technological advances demonstrate the commercial potential associated with this cryogenic technology.

7.3 Cryogenic hybrid systems

Traditionally, biogas upgrading technologies have focused on standalone techniques in order to remove CO2 and other contaminants from biogas. To overcome the possible drawbacks of standalone technologies, hybrid systems arise as a promising alternative for synergizing the best characteristics of biogas upgrading methods and/or exploring new paths for decreasing energy consumption/upgrading costs. Fig. 7.5 shows all the hybrid systems that have been studied to date for CO2 removal.

Figure 7.5 Hybrid systems for CO2 cryogenic removal. Adapted from Song C., Liu Q., Ji N., Deng S., Zhao J., Li Y., et al. 2018. Alternative pathways for efficient CO2 capture by hybrid processes—a review. Renew. Sustain. Energy Rev. 82, https://doi.org/10.1016/j.rser.2017.09.040.

As can be seen in Fig. 7.5, many combinations have been explored. The majority of which have been studied for CO2 capture and storage, but they could be potentially applied for biogas upgrading. In a recent study, these hybrid processes were classified into four categories, based on the main technology which was analyzed: absorption combinations, adsorption combinations, membrane combinations, and cryogenic combinations. In agreement with that study, the importance given by researchers in of investigation activities leads to sorting them as follows: (1) absorption combinations; (2) adsorption combinations; (3) membrane combinations; and (4) hydrate combinations (Song et al., 2018). Although cryogenic combinations are in the last position, there are many advantages which could open future research in this line. The main advantage of a low-energy CO2 capture biogas upgrading processes is the high CO2 product quality that can be obtained. Thus CO2 utilization and/or selling would be much easier than with the other techniques. Moreover, in the case of CO2 selling, no additional compression stages would be needed. Additionally, the use of cryogenic-based biogas upgrading systems avoids the utilization of solvents and hence minimizes any secondary pollution. In order to fulfill the authors’ concerns in this category of hybrid biogas upgrading processes, this section analyzes several promising options. Cryogenic hybrid systems are distinguished among cryogenic-absorption processes, cryogenic-adsorption processes, cryogenic-membrane processes, and cryogenic-hydrate processes. Some have been tested and studied for CO2 capture. Nevertheless, they could be potentially applied for biogas upgrading since the major contaminant to be removed from biogas streams is CO2. The most outstanding works in this area are analyzed in depth, showing the cryogenic hybrid proposals and the main outputs from the studies.

7.3.1 Cryogenic-absorption combination process

Low-temperature absorption processes consist of the utilization of a solvent at temperatures lower than usual. Under low-energy conditions, typical absorbents such as NaOH, KOH, or MEA do not show as high performance as they do at higher temperatures. Nevertheless, ammonia absorbs much more CO2 at low temperatures as low temperatures avoid ammonia evaporation. Moreover, ammonia is cheap, has huge availability, and low energy regeneration. Thus the employment of ammonia as a solvent for a range of temperatures of about 0°C– 10°C has been successfully tested by some research works. For example, a strict model was designed in order to validate and analyze a chilled ammonia process at low temperature for a 580 MWe supercritical coal-fired power plant. Several stripper pressure ranges (12.5–17.5 bar) were analyzed, with 12.5wt.% NH3 concentration. The cost investment of the plant was found to be 15.7% lower than the amines studies (Hanak et al., 2015). In the process depicted in Fig. 7.6 the sequence of operations to achieve CH4 and CO2 can be observed (Baena-Moreno et al., 2019f). In this process, a compression/cooling stage was previously needed to ensure that in the subsequent cryogenic stage the temperature was lower. In the cryogenic stage, the temperature of the biogas stream would be partially decreased. Nevertheless, the cryogenic temperatures for the direct separation of CO2–CH4 would not be achieved since the energy consumption would very high. Thus an intermediate temperature would be around −20°C. Afterwards, the biogas stream would be sent to the absorption/desorption stage in which ammonia would absorb the CO2 contained in the biogas. Biomethane would be obtained at the upper part of the absorption tower, whereas CO2-ammonia liquid current would be sent to the regeneration stage. Even though this process has not been analyzed for biogas upgrading, it could be an interesting future work to develop to find an equilibrium between the energy consumption of a cryogenic standalone process and the volatility of ammonia in absorption technology.

Figure 7.6 Cryogenic-absorption combination scheme for biogas upgrading. Adapted from Baena-Moreno F.M., Rodríguez-Galán M., Vega F., Vilches L.F., Navarrete B., Zhang Z., 2019f. Biogas upgrading by cryogenic techniques. Environ. Chem. Lett. 17, https://doi.org/10.1007/s10311-01900872-2.

7.3.2 Cryogenic-adsorption synergized process

Inasmuch that adsorption is a widely employed technique for biogas upgrading and CO2 capture (Vogtenhuber et al., 2018; Andriani et al., 2014), its potential combination with other technologies to overcome its disadvantages has been studied. A clear example of this is the process developed by Li Yuen Fong et al. (2016), where hybrid vacuum swing adsorption was synergized with a lowtemperature system. In agreement with this study, 1.40 MJ/t CO2 could be saved with an 88.9% CO2 recovery (Li Yuen Fong et al., 2016). The process proposed was intended to obtain a high-purity CO2 stream which could be sold to balance the economic performance of the process. Even though the main stages of this process are the adsorption and low-temperature (cryogenic) stages, temperature swing adsorption and membrane are also included. A tentative scheme of the proposed process by these authors adapted for biogas upgrading can be seen in Fig. 7.7. In this process first biogas upgrading would be introduced in a vacuum swing adsorption stage where a CH4-rich stream and a CO2 concentrated stream would be obtained. Afterwards, the CH4-rich stream would be sent to a membrane stage and temperature swing adsorption stage for dehydration where it would be further purified. On the other hand, the CO2 concentrated stream would be sent to a temperature swing adsorption stage for dehydrating and then to the cryogenic stage for its final purification and storage. Another potential cryogenic pressure–temperature swing adsorption method was analyzed by Moreira et al (2017). In this case, the process was applied to obtain LNG, and is similar in of technology for biogas upgrading. A CH4 recovery rate of 90.7% was obtained, with low impurities in comparison with standalone pressure —temperature swing adsorption (41.8 ppm of CO2). The total power consumption of the herein proposed process was 2.2 MW (Moreira et al., 2017).

Figure 7.7 Cryogenic-adsorption process. Adapted from Li Yuen Fong J.C., Anderson C.J., Xiao G., Webley P.A., Hoadley A.F.A., 2016. Multi-objective optimisation of a hybrid vacuum swing adsorption and low-temperature post-combustion CO2 capture. J. Clean. Prod. 111, https://doi.org/10.1016/j.jclepro.2015.08.033.

7.3.3 Potential combination of cryogenic and membrane processes

The combination of cryogenic processes with membrane technology is seen as a potential alternative to combine the advantages of both technologies. Indeed, the importance of this technology combination is enough for an entire section in this chapter (see Section 7.4). The process is schematized in Fig. 7.8 (Maqsood et al., 2017). Although there are numerous options for synergizing low-temperature processes with membrane technologies, Fig. 7.8 collects the most traditional one. The outstanding outcome of the typical cryogenic-membrane process is the lower cost in comparison with traditional MEA absorption, as was illustrated by Anantharaman et al. (2014). A 9% cost saving per ton of CO2 was concluded by these authors (Anantharaman et al., 2014). Further information about cryogenicmembrane processes can be seen in Section 7.4.

Figure 7.8 Process diagram for cryogenic-membrane biogas upgrading. Adapted from Maqsood K., Ali A., Shariff A.B.M., Ganguly S., 2017. Process intensification using mixed sequential and integrated hybrid cryogenic distillation network for purification of high CO2 natural gas. Chem. Eng. Res. Des. 117, https://doi.org/10.1016/j.cherd.2016.10.011.

7.3.4 Cryogenic-hydrate processes

Hydrate techniques for CO2 separation from biogas streams are a promising alternative that have led to some of the experts in this area studying this process in depth (Castellani et al., 2018; Filarsky et al., 2018; Lin et al., 2014). Hydrate techniques consist of subjecting biogas streams to high-pressure water currents. Thus the hydrate is formed and the CO2 is captured as part of it. Afterwards, in a subsequent stage, the hydrate and CO2 are separated. In a recent study, the carbon and energy footprints were found to be equal to 0.7081 kg CO2eq and 28.55 MJ/Nm³ of biomethane, respectively. The potential combination between hydrate technology and cryogenic separation arises as the needed process conditions for both technologies are low temperature and high pressure. The process schemed in Fig. 7.9 shows the typical setup for cryogenic-hydrate processes (Hart and Gnanendran, 2009). In the first stage, the temperature of the biogas stream would be subjected to −55°C, where a high portion of CO2 would be separated. The later stage would capture the remaining CO2 by hydrate technology. The working temperature of this second stage is about 1°C (Surovtseva et al., 2011). A high-purity CO2 stream for selling can be obtained by means of this combination. Additionally, utilization of this process would allow separation of the rest of the contaminants present in the biogas in the first stage.

Figure 7.9 Process diagram for cryogenic-hydrate biogas upgrading. Adapted from Hart A., Gnanendran N., 2009. Cryogenic CO2 capture in natural gas. Energy Proc. 1, https://doi.org/10.1016/j.egypro.2009.01.092.

7.4 Cryogenic-membrane processes

The combination of membranes and cryogenic techniques is one of the most promising strategies among the current initiatives in the field of biogas upgrading via cryogenics. Membranes have been developed in recent years for postcombustion CO2 capture applications (Merkel et al., 2010; Berstad et al., 2013). The potentialities of membranes are based on their simplicity, the use of environmental-friendly materials, and the energy savings for the CO2 separation stage (Hussain and Hägg, 2010; Song et al., 2019). Fig. 7.10 represents the twostage membrane process configuration in which 70% of the permeate is used as a feed stream for the second-stage membrane. In this configuration, 30% and 5% of the permeate is recirculated as a sweep in each stage for the first and second membranes, respectively (Song et al., 2019). Similar results could be obtained if this process was applied to biogas upgrading.

Figure 7.10 Two-stage membrane process configuration diagram flow. Adapted from Song C., Liu Q., Deng S., Li H., Kitamura Y., 2019. Cryogenic-based CO2 capture technologies: state-of-the-art developments and current challenges. Renew. Sustain. Energy Rev.101:265–278. https://doi.org/10.1016/j.rser.2018.11.018.

According to Belaissaoui et al. (2012a), multistage configuration using three membranes is the optimum strategy to achieve the highest efficiency under minimal investment cost (Fig. 7.11). Scholes et al. (2013) compared several membrane configurations in an exhaust gas from a cement kiln. The minimal energy consumption for CO2 separation was 1.2 MJ/kg CO2 using the threestage configuration, achieving CO2/N2 selectivity over 50. Further studies would be needed to estimate these results for biogas upgrading.

Figure 7.11 Three-stage membrane process configuration diagram flow. Adapted from Song C., Liu Q., Ji N., Deng S., Zhao J., Li Y., et al., 2017. Reducing the energy consumption of membrane-cryogenic hybrid CO2 capture by process optimization. Energy. 124, https://doi.org/10.1016/j.energy.2017.02.054.

However, the current development status for membrane materials and their low resilience under long-term operations make it unaffordable to achieve high CO2 separation targets using a single membrane separation technique (Mat and Lipscomb, 2019). Membranes also degrade in the presence of pollutants, particularly oxides from nitrogen and sulfur. High CO2 selectivity and high CO2 permeability are the most desirable properties, but the current development of conventional polymeric membrane exhibited a trade-off between the aforementioned properties (Merkel et al., 2010). In addition, membranes enhance the performance at subambient temperatures, achieving higher CO2/N2 selectivity levels despite lower CO2 permeability (Merkel et al., 2010). On the other hand, CO2 cryogenic condensation techniques can achieve a high CO2 purity liquid stream product from low CO2 concentrated treated gas but the energy requirements associated with cooling at the desired subambient temperature resulted in elevated operating costs (Song et al., 2017). To overcome these constrains, Belaissaoui et al. (2012b) and Merkel et al. (2010) proposed a novel hybrid membrane-cryogenic process in which a CO2 preconcentration stage using membranes supplied a higher CO2 flue gas for further CO2 separation based on cryogenics. The results taken from simulations done by Belaissaoui et al. revealed that the use of a CO2 concentrated flue gas ranging between 15–30%v/v in a hybrid cryogenic-membrane process resulted in energy penalties below 3 MJ/kg CO2 with CO2 purity above 89% and CO2 recovery ratios around 85% (Belaissaoui et al., 2012b), showing lower values than those required for a conventional MEA-based CO2 chemical absorption process— 4 MJ/kg CO2. Again, even though similar results could be obtained, further studies are needed to check the viability of this technology for biogas upgrading. Research activities into cryogenic-membrane technology have been carried out in recent years, focusing on three main topics: novel materials, cooling sources, and optimized configuration processes. In respect to novel materials for membranes operating at subambient temperatures, Mat and Lipscomb (2017)

proposed the use of materials that led to higher CO2 permeations with lower CO2/N2 selectivity. The use of this kind of materials can provide further reductions in the CO2 separation process. The reduction of the capital costs associated with high-permeability membranes should balance the inherent increase in the operating costs (Mat and Lipscomb, 2017). A novel Matrimid 5218 hollow-fiber membrane developed by Air Liquide has shown excellent performance operating at subambient temperatures. Based on the experiments performed in their works, Liu et al. demonstrated that the CO2/N2 selectivity of this novel membrane can be enhanced by between two and four times in comparison with ambient temperature operations (Liu et al., 2014). These experiments were set at temperatures below −20°C and the Matrimid 5218 hollow-fiber membrane reached up to 90.5% of CO2/N2 selectivity with minimal CO2 permeance losses with respect to the ambient temperature experiments (Hasse et al., 2014). Kumar et al. (2019) described a highly permeable ultra-thin skin CMS (ULT-CMS) membrane in which Matrimid was combined with polyvinylpyrrolidone and 6FDA/BPDA-DAM polymer. This novel membrane reported an up to two and three times increase in the CO2/N2 selectivity at subambient temperatures respect to ambient temperature experiments. All of these studies can be also applied for biogas upgrading. The cool energy supply is one of the most important constrains that hinders the deployment of this technology. Several authors have proposed the combination of membrane-cryogenic with natural gas or liquefied petroleum gas refrigeration systems (Garcia-Herrero et al., 2016). In this sense, the integration of a membrane-cryogenic technique with energy storage has been recently proposed (Fazlollahi et al., 2016; Safdarnejad et al., 2015). This novel concept is based on the production and storage of liquid natural gas during nonpeak periods in order to be used as a refrigerant to chill the treated gas in the CO2 separation stage. Modifications and novel configurations from the previously described membrane-cryogenic hybrid process proposed by Belaissaoui et al. (2012b) and Merkel et al. (2010) have been evaluated in recent years. Table 7.2 summarizes the most relevant modeling and technoeconomic analyses available in the literature. Substantial improvements in of energy consumption for the CO2 separation stage have been achieved compared to the reference case (Belaissaoui–Merkel), which almost doubled the specific energy consumption— 3 MJ/t CO2 (Belaissaoui et al., 2012b).

Table 7.2

Flue gas source

CO2 concentration (%)

Product purity (%)

Energy consumption (MJ/t CO2)

Gas combustion (coal)

13.9

95

0.94–1.08


15.0

99.8

1.70


13.5

98.3

1.22

Biogas

–

99.9 (CH4)

–


14.0

94.1

1.25


13.7

95

1.11

Mat and Lipscomb (2019) simulated a novel integration strategy for a membrane-cryogenic process for CO2 capture applications (Fig. 7.12).

Figure 7.12 The low-temperature membrane-cryogenic hybrid process configuration flow diagram. Adapted from Song C., Liu Q., Ji N., Deng S., Zhao J., Li Y., et al., 2017. Reducing the energy consumption of membranecryogenic hybrid CO2 capture by process optimization. Energy. 124, https://doi.org/10.1016/j.energy.2017.02.054.

The low-temperature membrane-cryogenic hybrid process consists of three sections. In the first section, the flue gas is compressed and then cooled by several heat exchangers in the low-temperature unit prior to being sent toward the membrane section. The compressed and cooled flue gas is introduced in the retentate flow side and generates the driving force to cause the CO2 permeation in the permeation flow side. The N2-enriched exhaust gas leaving the retentate flow side is then used to cool the entering gas, whereas the CO2 concentrated permeate gas is pumped and chilled before being introduced in the cryogenic separation unit. In this section, the CO2 is antisublimated into dry ice (solid), meanwhile the N2 es without phase change. All the cooled streams in this section are further heat integrated to provide the cooling energy required for the gas flow inlets. The increase in the CO2 concentration in the permeate gas fed to the cryogenic section reduces the volumetric flow rate entering the cryogenic section and increases the driving force in of CO2 partial pressure of the feed permeate gas. The combination of both effects leads to further reductions of the cooling energy penalties for the cryogenic section. The novel integration membrane-cryogenic process can lead to further improvements in of high CO2 purity, CO2 recovery, and energy consumption. According to Song et al., the novel integration process increased the CO2 purity of the permeate stream from 89% to 99.8%. The CO2 permeate stream temperature was increased from −110°C to−88.3°C, which implied a relevant reduction of the cooling energy requirements for the whole process. The optimized operation of this novel integration membrane-cryogenic process reduced the energy consumption by up to 1.7 MJ/t CO2, resulting in a 43% reduction compared to the baseline case (Song et al., 2017). The minimal energy consumption for the membrane-cryogenic hybrid process was obtained from work by Mat and Lipscomb (2019). In this work, the treated gas to be introduced in the cryogenic section was previously concentrated to

avoid excessive cooling of inert material such as N2. This work studied the assumption stated by Merkel et al. (2010) that indicated the cost of the membrane section dedicated to preconcentrating the treated gas by means of the use of the feed air to the boiler as a sweep of the membrane. The O2 concentration of the feed air to the boiler was kept above 18% to avoid combustion issues with the coal. According to Mat and Lipscomb (2019), a trade-off between the membrane area required for the preconcentration of the treated gas and the energy requirements for the CO2 separation should lead to optimization of the main operating parameters such as the pressure and the CO2/N2 selectivity. For feed pressures ranging between 2 and 4 bar, the optimal CO2/N2 selectivity was found to be in the range 55–65. Under the aforementioned operating conditions, the operating pressure and temperature in the condensation stage were set ranging between 15–18 bar and −50°C and −35°C, respectively. The combination of the operating parameters indicated above resulted in a specific energy consumption for the whole CO2 separation process ranging between 0.94 and 1.08 MJ/t CO2, a 67% reduction from the energy penalty compared with the conventional membrane-cryogenic configuration proposed by Belaissaoui et al. (2012b) and Merkel et al. (2010). All the operating conditions simulated in this work meet the targets in of CO2 purity (>90%) and CO2 recovery (>95%) (Mat and Lipscomb, 2019).

7.5 Full-scale experiences and technoeconomic studies

Cryogenics is currently considered an underdevelopment technology for biogas upgrading. This technology is capable to obtain high-purity biomethane (> 99%v/v CH4) with 98% recovery capacity (Angelidaki et al., 2018). According to the IEA and the European Biogas Association, cryogenic covers 1% of the total share of biogas upgrading in Europe (Hoyer et al., 2016). The low raw gas treatment capacity—below 500 Nm³/h—is responsible for the high CAPEX and OPEX costs which are hindering its deployment on a large scale (Prussi et al., 2019). Electricity consumption is identified as the major operational penalty. In this respect, the electricity consumption for cryogenic biogas applications was found to be around at 0.76 kWh/Nm³ which doubled the highest consumption reported for other mature biogas upgrading technologies (Angelidaki et al., 2018). Recent reports established the electricity consumption in 0.63 kWh/Nm³ but it is still far from economical sustainability of the process (Sahota et al., 2018). However, cryogenics can potentially provide high-purity liquid CO2 as a valuable subproduct which should make more economically feasible its application for biogas upgrading on a large scale (Hjuler and Aryal, 2017). Only a few full-scale experiences have been reported in the literature. The first commercial installation was erected in United States by Prometheus Energy. The plant capacity was 280 Nm³/h raw gas. The electricity requirements were extremely high, at approximately 1.54 kWh/Nm³ gas product (Andersson, 2009). In Europe, large-scale installations based on cryogenic technology were deployed by GtS. Two installations were built in Sweden in the period 2009–11. Their raw gas capacity ranged between 280–400 Nm³/h (Bauer et al., 2013). According to GtS, the electricity consumption can be substantially reduced in comparison with conventional values reported for pilot plant experiences and it was established at 0.45 kWh/Nm³. It should be noted that the licensor ensured CH4 losses below 0.5% and high-purity liquid CO2 as a subproduct. Despite the high performance exhibited from pilot plant tests, both projects were canceled in 2012 (Bauer et al., 2013). Currently, there is only one cryogenic biogas upgrading installation in operation

in Europe based on the Cryo Pur process (Berg, 2017). The plant, located in Northern Ireland, is the first farm-based bio-LNG in the world, producing biogas from animal and food waste treatment. It was erected in 2017 and has been fully operational since 2018. According to Berg, this plant produces 3 tpd of bio-LNG and 6 tpd of liquid CO2 from 340 Nm³/h raw gas. The licensor ensured that the capacity of the plant can be extended up to 2000 Nm³/h raw gas. In of energy requirements, the electricity consumption depends on the supplied conditions of the bio-LNG. It was reported from 0.6 to 0.7 kWh/Nm³ raw gas as the bio-LNG conditions vary from 15 bar (−120°C) to 2 bar (−160°C) (Berg, 2017). Finally, Cryostar has developed a liquefier system that can be adapted for biogas upgrading. This process, called StarLite LNG, is based on a nitrogen reverse Brayton closed loop process. According to the licensor, it can produce up to 200 tpd of liquid biogas with 0.6 kWh/kg (Berg, 2017). Concerning technoeconomic evaluations of biogas upgrading via cryogenic technology, only a few studies can be found. Estimations of cryogenic standalone technologies for CAPEX are around 394–960 €/kWh and for OPEX around 4.80–7.10 €/kWh (Baena-Moreno et al., 2019a). Tuinier et al (2011b) performed a technoeconomic evaluation of a cryogenic process, confirming an initial investment (total fixed capital) of 345 M$ and an operational cost of 52.8 $/t CO2 avoided (Tuinier et al., 2011b). These values are higher than for other commercial technologies such as membrane upgrading (Baena-Moreno et al., 2020d). Regarding hybrid cryogenic technologies, costs could be more competitive, at around 0.15 €/m³ of biomethane product (Scholz et al., 2013).

7.6 Comparison of documented technologies

Many technologies have been described in this chapter. For the sake of comparisons of the documented technologies, Table 7.3 includes the main advantages and disadvantages of each technology.

Table 7.3

Technology

Advantages

Cryogenic distillation

• High quality of the product streams (97.2% CH4; 99.3% CO2)

Cryogenic packed-bed technology

• Technology applied to a wide range of other gas separations • C

Cryogenic technology with absorption

• Utilization of a solvent at temperatures below the usual workin

Cryogenic technology with adsorption

• High potential to be combined with other technologies • Highly

Cryogenic technology with membranes

• The most promising strategy in the field of cryogenic biogas up

Cryogenic technology with hydrate technologies

• Promising alternative extensively studied by experts • High-pu

7.7 Conclusions and future perspectives

In this work cryogenic techniques as an alternative method for biogas upgrading have been analyzed in depth. As stated in the previous sections, cryogenic techniques provide high-purity CH4 and CO2 streams which allows to produce LBG and sell CO2. Nevertheless, the problem is the high energy consumption. Several works have been performed to reduce this energy consumption of standalone cryogenic techniques. The range of energy consumption is currently 0.5–1.5 kWh/Nm³ for cryogenic distillation and around 1.8 MJ/kg CO2 for B technology. The newly established paths for seeking novel alternatives which reduce energy consumption are the potential combination of cryogenic methods with other upgrading techniques. Thus over the last few years an evolution of absorption-cryogenic, adsorption-cryogenic, membrane-cryogenic, and hydratecryogenic methods has been carried out. Among the presented integrations, cryogenic-membrane technology is highlighted as the energy consumption can be considerably reduced without altering the purity of the final streams. The future works on cryogenic techniques should lead to scaling-up of the different configurations herein explained to achieve more realistic data which could help in the development of cryogenic technologies at industrial scales.

Appendix I Conversion factor for unit transformations

The use of each unit depends on the main objective that the installation is focused on. kWh/Nm³ is used for a conventional biogas upgrading process, whereas MJ/tCO2 is more specific from a CO2 capture point of view. The relationship between both units mainly depends on the ratio of CH4/CO2 content in the raw gas, which is different for each specific case. The conversion factor is indicated below:

Appendix II State forms for CO2 and CH4 as a function of temperature and pressure

Figure 7.AII1 State forms for CO2 as a function of temperature and pressure.

Figure 7.AII2 State forms for CH4 as a function of temperature and pressure.

Acknowledgments

This work was ed by the University of Seville through V PPIT-US. The authors acknowledge the assistance of Francisco Bueno Ramón for his technical help with the graphics presented in this work.

References

Ali et al., 2010 Ali G, Nitivattananon V, Molla NA, Hussain A. Selection of appropriate technology for solid waste management: a case of Thammasat hospital, Thailand. World Acad Sci Eng Technol. 2010;64:251–254. Ali et al., 2018 Ali A, Maqsood K, Shin LP, et al. Synthesis and mixed integer programming based optimization of cryogenic packed bed pipeline network for purification of natural gas. J Clean Prod. 2018;171:1–1690 https://doi.org/10.1016/j.jclepro.2017.10.060. Anantharaman et al., 2014 Anantharaman R, Berstad D, Roussanaly S. Technoeconomic performance of a hybrid membrane – liquefaction process for postcombustion CO2capture. Energy Proc. 2014;61:1244–1247 https://doi.org/10.1016/j.egypro.2014.11.1068. Andersson, 2009 Andersson K., 2009. Biogas till fordonsbränsle i form av flytande metan – Lämplig uppgraderingsteknik för Biogas Brålanda. Andriani et al., 2014 Andriani D, Wresta A, Atmaja TD, Saepudin A. A review on optimization production and upgrading biogas through CO2 removal using various techniques. Appl Biochem Biotechnol. 2014;172:1909–1928 https://doi.org/10.1007/s12010-013-0652-x. Angelidaki et al., 2018 Angelidaki I, Treu L, Tsapekos P, et al. Biogas upgrading and utilization: current status and perspectives. Biotechnol Adv. 2018;36:452– 466 https://doi.org/10.1016/j.biotechadv.2018.01.011. Baena-Moreno et al., 2018a Baena-Moreno FM, Rodríguez-galán M, Vega F, Alonso-fariñas B, Arenas LFV, Navarrete B. Carbon capture and utilization technologies: a literature review and recent advances. Energy Sour A Recover Util Environ Eff. 2018a;41:1–31 https://doi.org/10.1080/15567036.2018.1548518. Baena-Moreno et al., 2018b Baena-Moreno FM, Rodríguez-Galán M, Vega F,

Reina TR, Vilches LF, Navarrete B. Regeneration of sodium hydroxide from a biogas upgrading unit through the synthesis of precipitated calcium carbonate: an experimental influence study of reaction parameters. Processes. 2018b;6 https://doi.org/10.3390/pr6110205. Baena-Moreno et al., 2019a Baena-Moreno FM, Rodríguez-Galán M, Vega F, Vilches LF, Navarrete B. Review: recent advances in biogas purifying technologies. Int J Green Energy. 2019a;16:401–412 https://doi.org/10.1080/15435075.2019.1572610. Baena-Moreno et al., 2019b Baena-Moreno FM, Rodríguez-Galán M, Vega F, Reina TR, Vilches LF, Navarrete B. Synergizing carbon capture storage and utilization in a biogas upgrading lab-scale plant based on calcium chloride: influence of precipitation parameters. Sci Total Environ. 2019b;670 https://doi.org/10.1016/j.scitotenv.2019.03.204. Baena-Moreno et al., 2019c Baena-Moreno FM, Rodríguez-Galán M, Vega F, Ramirez-Reina T, Vilches L, Navarrete B. Understanding the influence of the alkaline cation K+ or Na+ in the regeneration efficiency of a biogas upgrading unit. Int J Energy Res. 2019c;43:1578–1585 https://doi.org/10.1002/er.4448. Baena-Moreno et al., 2019d Baena-Moreno FM, Rodríguez-Galán M, RamirezReina T, Zhang Z, Vilches L, Navarrete B. Understanding the effect of Ca and Mg ions from wastes in the solvent regeneration stage of a biogas upgrading unit. Sci Total Environ. 2019d;691:93–100 Available from: https://doi.org/10.1016/j.scitotenv.2019.07.135. Baena-Moreno et al., 2019e Baena-Moreno FM, Rodríguez-Galán M, Vega F, Reina TR, Vilches LF, Navarrete B. Converting CO2 from biogas and MgCl2 residues into valuable magnesium carbonate: a novel strategy for renewable energy production. Energy. 2019e;181 https://doi.org/10.1016/j.energy.2019.05.106. Baena-Moreno et al., 2019f Baena-Moreno FM, Rodríguez-Galán M, Vega F, Vilches LF, Navarrete B, Zhang Z. Biogas upgrading by cryogenic techniques. Environ Chem Lett. 2019f;17 https://doi.org/10.1007/s10311-019-00872-2. Baena-Moreno et al., 2020a Baena-Moreno FM, Pastor-Pérez L, Zhang Z, Reina TR. Stepping towards a low-carbon economy Formic acid from biogas as case of study. Appl Energy. 2020a;268 https://doi.org/10.1016/j.apenergy.2020.115033.

Baena-Moreno et al., 2020b Baena-Moreno FM, Pastor-Pérez L, Wang Q, Reina TR. Bio-methane and bio-methanol co-production from biogas: a profitability analysis to explore new sustainable chemical processes. J Clean Prod. 2020b;265 https://doi.org/10.1016/j.jclepro.2020.121909. Baena-Moreno et al., 2020c Baena-Moreno FM, Sebastia-Saez D, Wang Q, Reina TR. Is the production of biofuels and bio-chemicals always profitable? Co-production of biomethane and urea from biogas as case study. Energy Convers Manage. 2020c;220 https://doi.org/10.1016/j.enconman.2020.113058. Baena-Moreno et al., 2020d Baena-Moreno FM, le Saché E, Pastor-Pérez L, Reina TR. Membrane-based technologies for biogas upgrading: a review. Environ Chem Lett. 2020d;18 https://doi.org/10.1007/s10311-020-01036-3. Bauer et al., 2013 Bauer F., Hulteberg C., Persson T., Tamm D., 2013. Biogas upgrading – review of commercial technologies. Swedish Gas Technol Centre, SGC 2013:82. doi: SGC Rapport:270. Belaissaoui et al., 2012a Belaissaoui B, Le Moullec Y, Willson D, Favre E. Hybrid membrane cryogenic process for post-combustion CO2 capture. J Memb Sci. 2012a;415-416 https://doi.org/10.1016/j.memsci.2012.05.029. Belaissaoui et al., 2012b Belaissaoui B, Willson D, Favre E. Membrane gas separations and post-combustion carbon dioxide capture: parametric sensitivity and process integration strategies. Chem Eng J. 2012b;211-212 https://doi.org/10.1016/j.cej.2012.09.012. Benjaminsson et al., 2013 Benjaminsson G., Benjaminsson J., Rudberg R. 2013. Power-to-gas – a technical review. SGC Rapp 2013. doi: SGC Rapport:284. Berg, 2017 Berg B., 2017. Efficient liquid biomethane production with cryogenic upgrading. Gas Energy. Berstad et al., 2013 Berstad D, Anantharaman R, Nekså P. Low-temperature CO2 capture technologies – applications and potential. Int J Refrig. 2013;36 https://doi.org/10.1016/j.ijrefrig.2013.03.017. Castellani et al., 2018 Castellani B, Rinaldi S, Bonamente E, Nicolini A, Rossi F, Cotana F. Carbon and energy footprint of the hydrate-based biogas upgrading process integrated with CO2 valorization. Sci Total Environ. 2018;615

https://doi.org/10.1016/j.scitotenv.2017.09.254. DiMaria et al., 2015 DiMaria PC, Dutta A, Mahmud S. Syngas purification in cryogenic packed beds using a one-dimensional pseudo-homogenous model. Energy Fuels. 2015;29 https://doi.org/10.1021/acs.energyfuels.5b00624. Energy GF Energy GF. Biogas upgrading. n.d. <www.gtsbv.com/overgts/biogas-opwaardering> (accessed 16.11.19). Fazlollahi et al., 2016 Fazlollahi F, Bown A, Ebrahimzadeh E, Baxter LL. Transient natural gas liquefaction and its application to CCC-ES (energy storage with cryogenic carbon capture™). Energy. 2016;103 https://doi.org/10.1016/j.energy.2016.02.109. Filarsky et al., 2018 Filarsky F, Schmuck C, Schultz HJ. Development of a biogas production and purification process using promoted gas hydrate formation—a feasibility study. Chem Eng Res Des. 2018;134 https://doi.org/10.1016/j.cherd.2018.04.009. Garcia-Herrero et al., 2016 Garcia-Herrero I, Cuéllar-Franca RM, EnríquezGutiérrez VM, Alvarez-Guerra M, Irabien A, Azapagic A. Environmental assessment of dimethyl carbonate production: comparison of a novel electrosynthesis route utilizing CO2 with a commercial oxidative carbonylation process. ACS Sustain Chem Eng. 2016;4:2088–2097 https://doi.org/10.1021/acssuschemeng.5b01515. Hanak et al., 2015 Hanak DP, Biliyok C, Manovic V. Rate-based model development, validation and analysis of chilled ammonia process as an alternative CO2 capture technology for coal-fired power plants. Int J Greenh Gas Control. 2015;34 https://doi.org/10.1016/j.ijggc.2014.12.013. Hart and Gnanendran, 2009 Hart A, Gnanendran N. Cryogenic CO2 capture in natural gas. Energy Proc. 2009;1 https://doi.org/10.1016/j.egypro.2009.01.092. Hasse et al., 2014 Hasse D, Ma J, Kulkarni S, et al. CO2 capture by cold membrane operation. Energy Proc. 2014;63 https://doi.org/10.1016/j.egypro.2014.11.019. Hjuler and Aryal, 2017 Hjuler K., Aryal N., 2017. Review of biogas upgrading futuregas project, WP1. Dansk Gasteknisk Center. doi: 978-87-7795-403-0.

Horschig et al., 2019 Horschig T, Welfle A, Billig E, Thrän D. From Paris agreement to business cases for upgraded biogas: analysis of potential market uptake for biomethane plants in using biogenic carbon capture and utilization technologies. Biomass Bioenergy. 2019;120 https://doi.org/10.1016/j.biombioe.2018.11.022. Hoyer et al., 2016 Hoyer, K., Hulteberg, C., Svensson, M., Jernberg, J., Nörregård Ö., 2016, Biogas upgrading – technical review. Hussain and Hägg, 2010 Hussain A, Hägg MB. A feasibility study of CO2 capture from flue gas by a facilitated transport membrane. J Memb Sci. 2010;359 https://doi.org/10.1016/j.memsci.2009.11.035. Kumar et al., 2019 Kumar R, Zhang C, Itta AK, Koros WJ. Highly permeable carbon molecular sieve membranes for efficient CO2/N2 separation at ambient and subambient temperatures. J Memb Sci. 2019;583 https://doi.org/10.1016/j.memsci.2019.04.033. Li Yuen Fong et al., 2016 Li Yuen Fong JC, Anderson CJ, Xiao G, Webley PA, Hoadley AFA. Multi-objective optimisation of a hybrid vacuum swing adsorption and low-temperature post-combustion CO2 capture. J Clean Prod. 2016;111 https://doi.org/10.1016/j.jclepro.2015.08.033. Lin et al., 2014 Lin H, He Z, Sun Z, et al. CO2-selective membranes for hydrogen production and CO2 capture – part I: membrane development. J Memb Sci. 2014;457 https://doi.org/10.1016/j.memsci.2014.01.020. Liu et al., 2014 Liu L, Sanders ES, Kulkarni SS, Hasse DJ, Koros WJ. Subambient temperature flue gas carbon dioxide capture via Matrimid® hollow fiber membranes. J Memb Sci. 2014;465 https://doi.org/10.1016/j.memsci.2014.03.060. Maqsood et al., 2017 Maqsood K, Ali A, Shariff ABM, Ganguly S. Process intensification using mixed sequential and integrated hybrid cryogenic distillation network for purification of high CO2 natural gas. Chem Eng Res Des. 2017;117 https://doi.org/10.1016/j.cherd.2016.10.011. Mat and Lipscomb, 2017 Mat NC, Lipscomb GG. Membrane process optimization for carbon capture. Int J Greenh Gas Control. 2017;62 https://doi.org/10.1016/j.ijggc.2017.04.002.

Mat and Lipscomb, 2019 Mat NC, Lipscomb GG. Global sensitivity analysis for hybrid membrane-cryogenic post combustion carbon capture process. Int J Greenh Gas Control. 2019; https://doi.org/10.1016/j.ijggc.2018.12.023. Merkel et al., 2010 Merkel TC, Lin H, Wei X, Baker R. Power plant postcombustion carbon dioxide capture: an opportunity for membranes. J Memb Sci. 2010;359 https://doi.org/10.1016/j.memsci.2009.10.041. Moreira et al., 2017 Moreira MA, Ribeiro AM, Ferreira AFP, Rodrigues AE. Cryogenic pressure temperature swing adsorption process for natural gas upgrade. Sep Purif Technol. 2017;173 https://doi.org/10.1016/j.seppur.2016.09.044. Nilsson, 2001 Nilsson M.M.L.A., 2001. Livscykelinventering för biogas som fordons- bränsle. Pellegrini et al., 2018 Pellegrini LA, De Guido G, Langé S. Biogas to liquefied biomethane via cryogenic upgrading technologies. Renew Energy. 2018;124 https://doi.org/10.1016/j.renene.2017.08.007. Persson Owe Jönsson Arthur Wellinger et al., 2006 Persson Owe Jönsson Arthur Wellinger M., Margareta Persson A., Jönsson O., Wellinger A., Rudolf Braun A., Jens Bo Holm-Nielsen D., 2006. Biogas upgrading to vehicle fuel standards and grid injection. Prussi et al., 2019 Prussi M, Padella M, Conton M, Postma ED, Lonza L. Review of technologies for biomethane production and assessment of Eu transport share in 2030. J Clean Prod. 2019;222 https://doi.org/10.1016/j.jclepro.2019.02.271. Safdarnejad et al., 2015 Safdarnejad SM, Hedengren JD, Baxter LL. Plant-level dynamic optimization of cryogenic carbon capture with conventional and renewable power sources. Appl Energy. 2015;149 https://doi.org/10.1016/j.apenergy.2015.03.100. Sahota et al., 2018 Sahota S, Shah G, Ghosh P, et al. Review of trends in biogas upgradation technologies and future perspectives. Bioresour Technol Rep. 2018;1 https://doi.org/10.1016/j.biteb.2018.01.002. Scholes et al., 2013 Scholes CA, Ho MT, Wiley DE, Stevens GW, Kentish SE.

Cost competitive membrane-cryogenic post-combustion carbon capture. Int J Greenh Gas Control. 2013;17 https://doi.org/10.1016/j.ijggc.2013.05.017. Scholz et al., 2013 Scholz M, Frank B, Stockmeier F, Falß S, Wessling M. Techno-economic analysis of hybrid processes for biogas upgrading. Ind Eng Chem Res. 2013;52 https://doi.org/10.1021/ie402660s. Shafiee et al., 2017 Shafiee A, Nomvar M, Liu Z, Abbas A. Automated process synthesis for optimal flowsheet design of a hybrid membrane cryogenic carbon capture process. J Clean Prod. 2017;150 https://doi.org/10.1016/j.jclepro.2017.02.151. Song et al., 2017 Song C, Liu Q, Ji N, et al. Reducing the energy consumption of membrane-cryogenic hybrid CO2 capture by process optimization. Energy. 2017;124 https://doi.org/10.1016/j.energy.2017.02.054. Song et al., 2018 Song C, Liu Q, Ji N, et al. Alternative pathways for efficient CO2 capture by hybrid processes—a review. Renew Sustain Energy Rev. 2018;82 https://doi.org/10.1016/j.rser.2017.09.040. Song et al., 2019 Song C, Liu Q, Deng S, Li H, Kitamura Y. Cryogenic-based CO2 capture technologies: state-of-the-art developments and current challenges. Renew Sustain Energy Rev. 2019;101:265–278 https://doi.org/10.1016/j.rser.2018.11.018. Sun et al., 2015 Sun Q, Li H, Yan J, Liu L, Yu Z, Yu X. Selection of appropriate biogas upgrading technology - a review of biogas cleaning, upgrading and utilisation. Renew Sustain Energy Rev. 2015;51:521–532 https://doi.org/10.1016/j.rser.2015.06.029. Surovtseva et al., 2011 Surovtseva D, Amin R, Barifcani A. Design and operation of pilot plant for CO2 capture from IGCC flue gases by combined cryogenic and hydrate method. Chem Eng Res Des. 2011;89 https://doi.org/10.1016/j.cherd.2010.08.016. Tuinier et al., 2010 Tuinier MJ, van Sint Annaland M, Kramer GJ, Kuipers JAM. Cryogenic CO2 capture using dynamically operated packed beds. Chem Eng Sci. 2010;65 https://doi.org/10.1016/j.ces.2009.01.055. Tuinier et al., 2011a Tuinier MJ, Van Sint Annaland M, Kuipers JAM. A novel

process for cryogenic CO2 capture using dynamically operated packed beds—an experimental and numerical study. Int J Greenh Gas Control. 2011a;5 https://doi.org/10.1016/j.ijggc.2010.11.011. Tuinier et al., 2011b Tuinier MJ, Hamers HP, Van Sint Annaland M. Technoeconomic evaluation of cryogenic CO2 capture—a comparison with absorption and membrane technology. Int J Greenh Gas Control. 2011b;5 https://doi.org/10.1016/j.ijggc.2011.08.013. Vogtenhuber et al., 2018 Vogtenhuber H, Hofmann R, Helminger F, Schöny G. Process simulation of an efficient temperature swing adsorption concept for biogas upgrading. Energy. 2018;162 https://doi.org/10.1016/j.energy.2018.07.193. Yousef et al., 2017 Yousef AM, Eldrainy YA, El-Maghlany WM, Attia A. Biogas upgrading process via low-temperature CO2 liquefaction and separation. J Nat Gas Sci Eng. 2017;45 https://doi.org/10.1016/j.jngse.2017.07.001. Yousef et al., 2018 Yousef AM, El-Maghlany WM, Eldrainy YA, Attia A. New approach for biogas purification using cryogenic separation and distillation process for CO2 capture. Energy. 2018;156 https://doi.org/10.1016/j.energy.2018.05.106. Yousef et al., 2019 Yousef AM, El-Maghlany WM, Eldrainy YA, Attia A. Upgrading biogas to biomethane and liquid CO2: a novel cryogenic process. Fuel. 2019;251 https://doi.org/10.1016/j.fuel.2019.03.127. Zhang and Lior, 2006 Zhang N, Lior N. A novel near-zero CO2 emission thermal cycle with LNG cryogenic exergy utilization. Energy. 2006;31 https://doi.org/10.1016/j.energy.2005.05.006. Zhang and Lior, 2008 Zhang N, Lior N. Two novel oxy-fuel power cycles integrated with natural gas reforming and CO2 capture. Energy. 2008;33 https://doi.org/10.1016/j.energy.2007.09.006. Zhang et al., 2010 Zhang N, Lior N, Liu M, Han W. COOLCEP (cool clean efficient power): a novel CO2-capturing oxy-fuel power system with LNG (liquefied natural gas) coldness energy utilization. Energy. 2010;35 https://doi.org/10.1016/j.energy.2009.04.002.

Zhang et al., 2014 Zhang X, Singh B, He X, Gundersen T, Deng L, Zhang S. Post-combustion carbon capture technologies: energetic analysis and life cycle assessment. Int J Greenh Gas Control. 2014;27 https://doi.org/10.1016/j.ijggc.2014.06.016. Zhang et al., 2020 Zhang N, Pan Z, Zhang Z, et al. CO2 capture from coalbed methane using membranes: a review. Environ Chem Lett. 2020;18 https://doi.org/10.1007/s10311-019-00919-4. Zhao et al., 2014 Zhao L, Primabudi E, Stolten D. Investigation of a hybrid system for post-combustion capture. Energy Proc 2014; https://doi.org/10.1016/j.egypro.2014.11.183.

Chapter 8

Power-to-gas for methanation

Anirudh Bhanu Teja Nelabhotla¹, Deepak Pant² and Carlos Dinamarca¹, ¹1Department of Process Energy and Environmental Technology, University of South-Eastern Norway, Porsgrunn, Norway, ²2Separation and Conversion Technology, Flemish Institute for Technological Research (VITO), Mol, Belgium

Abstract

Power-to-gas (PtG) is a technology in development in the field of renewable energy (RE) and sustainable energy management. RE such as solar and wind are unreliable sources of energy due to their dependence on nature, which is highly unpredictable. This makes electricity production via the aforementioned sources largely intermittent and fluctuating. Large-scale mechanisms for energy storage to mitigate output fluctuations are complex, as several technical solutions (e.g., hydroelectricity storage and batteries) are often required, with integration, coordination, and planning still in the early stages of study and development. PtG can convert renewable electricity into a flexible fuel, that is, gas, that can be easily stored, transported, and converted to other valuable chemicals, diversifying the end use of RE. PtG integrates renewable electricity sources with a storage system while consuming other pollutants such as CO2 and waste. Methane, the end product in this case, is a highly efficient gaseous fuel and is carbon neutral when generated via PtG.

Keywords

Power-to-gas; methanation; electrolyzers; biocathode; reactor design

Chapter outline

Outline


8.2 Electrocatalytic methanation 189

8.2.1 Alkaline electrolyzers 190 8.2.2 Polymer electrolyte membrane electrolyzers 193 8.2.3 Solid oxide electrolyzers 196 8.2.4 Fixed-bed methanation reactors 198 8.2.5 Fluidized bed methanation reactors 200 8.2.6 Three-phase reactor 202 8.2.7 Micro(channel) reactors 203

8.3 Bioelectrochemical methanation 205

8.3.1 Direct electron transfer 207 8.3.2 Biocathodes 209 8.3.3 Reactor configurations 210

8.4 Challenges and future prospects 211

References 213

8.1 Introduction

Power-to-gas (PtG) is technology in development in the field of renewable energy (RE) and sustainable energy management. RE such as solar and wind are unreliable sources of energy due to their dependence on nature, which is highly unpredictable (Hashimoto et al., 2014). This makes electricity production via the aforementioned sources largely intermittent and fluctuating (del Pilar Anzola Rojas et al., 2018; Anzola-Rojas et al., 2018). Large-scale mechanisms for energy storage to mitigate output fluctuations are complex as several technical solutions (e.g., hydroelectricity storage and batteries) are often required, the integration, coordination, and planning of which are still in the early stage of study and development (Yekini Suberu et al., 2014; Hemmati et al., 2017). PtG can convert renewable electricity into a flexible fuel, that is, gas, that can be easily stored, transported, and converted to other valuable chemicals, diversifying the end use of RE (Breyer et al., 2015). PtG integrates renewable electricity sources with a storage system while consuming pollutants such as CO2 (Pestalozzi et al., 2019). Methane, the end product in this case, is a highly efficient gaseous fuel and is carbon neutral when generated via PtG (Geppert et al., 2019). While there have been many technologies explored for storing electricity, some methods have certain advantages over others. Two of the main criteria evaluated for such a comparison are the storage capacity versus discharge time. It is important that the best approach for storing electricity has a combination of large storage capacity and at the same time maintaining high discharge times. The discharge time of an energy source corresponds to the time the source can maintain its energy output. Table 8.1 compares the discharge time and storage capacities of some of the prominent methods to store electricity (Specht et al., 2010). Carbon-based liquid, gaseous fuels, hydrogen, and other chemical secondary energy carriers are able to provide such a combination of energy storage and utilization mechanism (Schaaf et al., 2014).

Table 8.1

Method

Discharge time

Storage capacity

Supercapacitors and superconducting magnetic energy storage

Low

Low

Pumped storage in power plants

Medium

Medium

Batteries

Low

Medium

Compressed air reservoirs

Medium

Medium

PtG (hydrogen, methane, and synthetic natural gas)

High

Large

PtG technology provides an opportunity to reintroduce carbon dioxide into the energy cycle by reducing it to various chemicals or gases. This in turn s production of carbon-based energy-rich products from electricity and carbon dioxide. CO2 can be sourced from polluting industries (Uhm and Kim, 2014), waste treatment facilities (Biesemans, 2016), direct air capture (Peters et al., 2019), and other carbon capture and storage (CCS) technologies. The current CCS research and development proposes to store the captured CO2 in pressurized cylinders in the seabed (Braathen et al., 2019). This solution, although capable of reducing the impact of CO2 on climate change temporarily, does not motivate to decarbonize or reduce the carbon footprint of the energy production process and energy-intensive industries. Such an incentive can be created when carbon dioxide, now an environmental harmful end product, can be turned into a valuable resource (Antenucci and Sansavini, 2019). PtG integrates the RE sector with carbon capture technology developers and has the potential to close the carbon cycle loop in the industrial sector. The main advantages of the production and utilization of a gaseous fuel such as methane from PtG are storage and transportation. An energy resource must be accessible to all citizens no matter where they live and how much they use. Transporting electricity to remote locations can be difficult due to the expensive and heavy infrastructure involved rather than transporting a gaseous fuel. Moreover, not all peoples' needs can be solved by electricity, especially in poorer societies and remote localities where they cannot afford expensive and modern equipment that runs on electricity. Many countries around the world such as , Denmark, the United States, and Canada have an existing natural gas grid that is compatible for transporting methane from PtG (Eveloy and Gebreegziabher, 2018). An exponential increase in PtG installed capacity is predicted as the electrolysis and methanation costs are estimated to fall by up to 75% under 500 €/kWel until 2050 (Thema et al., 2019). This will further encourage other countries to adopt PtG and contribute to the predicted growth of the technology. A gaseous fuel is more flexible in of an energy resource as it could be used for generating heat, cooking food, and most importantly can be used as a transport fuel, which is currently oil intensive. Carbon dioxide reduction to methane can broadly be achieved via two methods: (1) chemical methanation and (2) bioelectrochemical methanation. Chemical

methanation at present is marginally advanced as compared to bioelectrochemical methanation in of research and large-scale production prototypes. Chemical methanation involves high-temperature and high-pressure reactors using medium to highly expensive chemical and metal catalysts as opposed to bioelectrochemical methanation that relies on ambient temperatures and pressures and microorganisms as catalysts. Bioelectrochemical methanation has not been demonstrated in large scale apart from a very few pilot studies and is still under development. However, both technologies are yet to break into the mainstream commercial market. Some of the primary distinctions between chemical and bioelectrochemical methanations are mentioned in Table 8.2, followed by descriptions of specific processes in both production mechanisms.

Table 8.2

Parameter

Chemical

Bioelectrochemical

Temperature

200°C–800°C

20°C–50°C

Pressure

20–100 bars

1–3 bar

Catalysts

Expensive nobel metals

Microorganisms

Reactor footprint

Small

Small

Operational flexibility

Medium–rigid

Flexible

Capital expenditure

High

Medium

Operational expenditure

High

Medium

Production efficiencies

High

Medium

8.2 Electrocatalytic methanation

An electrochemical methane production that involves electrolysis followed by use of metal catalysts to reduce CO2 can be termed as electrocatalytic methanation. The methanation technology was first developed as a cleaning and purification process for the production of other chemicals such as hydrogen gas (Xu et al., 2006) and ammonia (Appl, 1999). These methods have now found multiple applications with methane production being instrumental to storing excess RE. Electrocatalytic methanation is the primary method to convert electricity to methane and is carried out in two steps, namely, electrolysis [Eqs. (8.1) and (8.2)] and methanation [(Eq. (8.3)]. Electrolysis is carried out in electrolyzers that have been approached in many ways such as alkaline water electrolysis (AWE), polymer electrolyte fuel cells (PEFCs), proton exchange membrane (PEM), and solid oxide electrolysis cells (SOECs). This is followed by methanation where CO2, CO or a mixture of both is combined with hydrogen in one of many variations of fixed-bed, fluidized bed, three-phase, and other reactor designs being developed. Electrolysis reactions (Rabaey and Rozendal, 2010) include the following: Anode reaction

(8.1)

Cathode reaction

(8.2)

Methanation reactions (Schaaf et al., 2014; Domigall, 2016)

(8.3)

(8.4)

8.2.1 Alkaline electrolyzers

8.2.1.1 Definition and concept

AWE is a mature and safe technology and is currently used in many industrial applications. One of the main advantages of using AWE over other technologies is its ability to scale-up to megawatt range production capacities (Ursúa et al., 2012). However, the process is very energy demanding with high capital investment, operation, and maintenance costs (Manabe et al., 2013). AWE is based on an alkaline electrolyte such as sodium hydroxide (NaOH) or potassium hydroxide (KOH) that is added to the water, typically in concentrations of about 20–40 wt.%. AWE is operated at approximately 70–90°C and at a maximum pressure of about 30 bar (Rashid et al., 2015). The hydroxide ion is produced at the cathode by splitting water which carries the charge across a

membrane/diaphragm and is discharged at the anode by producing oxygen. The chemical Eqs. (8.1) and (8.2) are modified to Eqs. (8.5)– (8.9) to carry out AWE, where Eq. (8.5) is the base cathode reaction that occurs in steps of either (8.6) and (8.7) called the Tafel reaction, or (8.6) and (8.8) called the Heyrovsky reaction. Eq. (8.9) is the modified anode reaction where four moles of hydroxide ion oxidize to water and oxygen, releasing 4 mol electrons (Marini et al., 2012). AWE reactions include the following: Cathode reactions

(8.5)

(8.6)

Tafel reaction

(8.7)

Or Heyrovsky reaction

(8.8)

Anode reaction

(8.9)

Eq. (8.6) indicates the role of electrodes in electrolytic hydrogen production by adsorbing the proton at site “M” and releasing the hydroxide ion which is the electron carrier. The Tafel and Heyrovsky reactions show the second step in the reaction where hydrogen gas is produced, and the electrode site is made available for the next round of adsorption. The adsorption of protons on the electrode site is done with the help of a hydrogen bond that is not as strong as it is not reversible and is able to form hydrogen combining with another proton that is either attached to a nearby electrode site (Tafel reaction) or another water molecule in the alkaline electrolyte (Heyrovsky reaction).

8.2.1.2 Reactor configurations

AWE has one of the simplest reactor configurations that consists of a cathode, anode, electrolyte, and a two-chamber reactor-cell separated by a diaphragm (Fig. 8.1). The electrodes or electrocatalysts are the defining components of the electrolysis process; hydrogen is produced on the cathode, while oxygen is formed on the anode. The electrode material determines the efficiency of the electrolysis as it reduces the activation energy (Ea) and improves the reaction kinetics of both the hydrogen and oxygen formation reaction pathways. Electrolytes such as NaOH and KOH are commonly used for ionic activation of the electrolysis reaction. However, these alkaline solutions are highly corrosive and require the addition of other chemicals and ionic liquids that neutralize the

corrosion effect. Some of these chemicals include Na2MoO4, Na2WO4 ethylenediamine-based metal chloride complex ([M(en)3]Clx, M=Co, Ni, etc.), ethylene diamine cobalt (III) chloride complex ([Co(en)3]Cl3) or trimethylenediamine cobalt(III) chloride complex ([Co(tn)3]Cl3), Na2MoO4, and [Ni (en)3]Cl2 (Tasic et al., 2011; Nikolic et al., 2010; Marceta Kaninski et al., 2011; Maksic et al., 2011; Stojić et al., 2003). The addition of these chemicals protects the electrode material from corrosion and degradation, providing a longer operating life and thus reaping economic benefits.

Figure 8.1 Schematic representation of alkaline water electrolyzers (Khatib et al., 2019).

These components are placed in an electrolysis cell where a diaphragm is placed at the center of the cell, creating two chambers. The chambers are designed to hold two electrodes separately, either at opposite ends or at the center of each of the two chambers. The diaphragm is a microporous material used to separate the gas products at each electrode terminal and facilitate the transport of hydroxyl ions. To yield high cell efficiency the diaphragm must possess high water permeation, withstand corrosion in strongly alkaline media, and have high ionic conductivity (Coutanceau et al., 2018). Previously, asbestos was used to manufacture diaphragms, but since it is banned now, other alternatives are used such as composite of potassium titanate (K2TiO3) fibers and polytetrafluoroethylene (PTFE), polyphenylene sulfide, PTFE (as felt and as woven), and polysulfone (Rosa et al., 1995). Many electrode materials and reactor configurations have been studied with little modification over the conventional design. This conventional design of the electrolysis cell, however, is inefficient in of ohmic losses due to the distance between electrodes.

8.2.1.3 Recent developments

In 2011, a new technique called the “zero-gap cell” was developed on a alkaline electrolyzer to overcome the problem of ohmic losses during the electron and mass transfer (Li et al., 2011). The conventional solid electrode rods and diaphragm setup were modified to a porous cathode and anode attached to either side of the diaphragm that acts as a gas separator with no electrolyte in between. These modifications to the reactor design significantly improved the current density by eradicating the losses caused by the formation of bubbles and thereby improved hydrogen production efficiencies close to that of PEM electrolyzers (Phillips and Dunnill, 2016). Further improvement in the technology was establish by Marini et al., with the introduction of gas diffusion electrodes (GDEs) into AWE (Marini et al., 2012). The new configuration places a

recirculating electrolyte sandwiched between the cathode and anode assemblies on either side. The GDE assemblies are made of Ni-net porous electrode, gas diffusion layer (GDL), active layer, Celgard, or ZrO2 separator (Marini et al., 2012). Both the anode and cathode assemblies are open to their respective gas compartments that collect and release the O2 and H2, respectively, diffused through the GDLs. The authors conclude that such a system is as efficient as a zero-gap system, while being suitable for industrial production due to its simpler system requirements.

8.2.2 Polymer electrolyte membrane electrolyzers

8.2.2.1 Design and concept

The concept of PEM was developed to overcome the negative aspects of AWE, such as the lower current densities and lower gas purity. This was done with the use of a solid polymer electrolyte and demonstrated for the first time using solid sulfonated polystyrene membrane (Fig. 8.2) (Grubb, 1959a,b). Therefore the process is called PEM water electrolysis or solid polymer electrolysis. The polymer membrane also acts as a gas separator by letting protons cross from the anode to the cathode. Thus the electrolysis process is named proton exchange membrane. Eqs. (8.1) and (8.2) in Section 8.2, describe the anode and cathode reactions, respectively. PEM electrolysis occur at temperatures ranging between 20°C–100°C and at a maximum reactor pressure of 30 bars. The high-efficiency PEM can generate a maximum current density of up to 2 A/cm², that is, 2–5 times more than AWE (Michishita et al., 2008; Takenaka, 1991).

Figure 8.2 Schematic representation of polymer electrolyte membrane electrolyzers (Khatib et al., 2019).

Other advantages of PEM include low or no chemical usage, compact systems, low carbon footprint, high hydrogen purity and proton conductivity implying low gas crossover, lower power usage, high operation pressure, easy maintenance, control over power fluctuations, thus dynamic operation, and higher safety level (Grigoriev and Porembsky, 2006). However, scaling up PEM was difficult due to the higher costs of membrane and other reactor components, allowing the stack capacities to be below the MW range (Carmo et al., 2013). The acidic environment of the reactor leads to corrosion of the expensive components such as electrodes and membrane. This also makes the durability of the electrolysis system low for longer operational periods (Barbir, 2005). Continuous research for the development of PEM electrolysis is underway with innovative and economical ways to manufacture electrolyte membranes, catalysts, and current collecting materials.


The main element of a PEM reactor is the membrane electrode assembly (MEA), where the electrodes, gas diff, gasket, bipolar plates, and interconnector are coated on each side of the membrane and placed in the reactor (Rashid et al., 2015). The polymer electrolyte membrane (Nafion, Flemion, Fumapem, and Aciplex) provides high proton conductivity and other advantages, as described previously. The low membrane thickness (20–300 μm) explains in part many of the benefits of the solid polymer electrolyte (Carmo et al., 2013). The membranes have lifetimes of approximately 10,000 h, with high proton conductivities of about 0.1 S/cm. The MEA is assembled most commonly by hot pressing the electrode material and current collector on to the membrane surface. The electrocatalyst is coated on to the electrode materials at a cathode load of up to 2 mg cm² and for an

anode up to 6 mg cm². The most common electrocatalyst materials are platinum for the cathode, and iridium or rhodium for the anode. The rare and expensive materials used for electrocatalysts explain the high price for MEA and thereby PEM. This is the challenge in scaling up the PEM technology, as the size of the reactor increases the required electrode surface area, leading to high capital and maintenance costs. Many studies have focused on alternative cheaper electrocatalysts for both oxygen evolution reaction (OER) and hydrogen evolution reaction (HER) (Furuya and Motoo, 1979; Buckley, 1974). Very few researchers have made successful attempts at bringing the costs down even by small margins. Still PEM is the most preferred form of electrolysis technology to yield H2-gas because of its efficiency and compact build. Diluting these metal anodes with their oxides such as IrO2 and RhO2 is considered to increase the stability of the electrode material against corrosion, however, the results were not consistent (Mamaca et al., 2012; Fuentes et al., 2010). Similarly, a platinum cathode combined with carbon or Cu nanoparticles has also been tested, showing no significant improvement (Raoof et al., 2010). More recent research activities were focused on electrode metal design using binary, ternary, or quaternary alloys for increased stability and increased lifetime.


Many nanoalloys have been developed with the primary aim of reducing the HER- related costs, some examples are A-Ni-C, Mo2C/CNTs, Ni2P/CNTs, Codoped FeS2/CNTs, WO2/C nanowires, and CoFe nanoalloys encapsulated in Ndoped graphene (Shiva Kumar and Himabindu, 2019). Recent studies showed improved HER performance with increased current density using synthesized RuS2@MoS2 nanoparticle–nanosheet alloy catalyst (Sarno and Ponticorvo, 2019). Palladium has been associated with many recent innovative cathode applications due to its platinum-like properties. Additionally, plladium is cheap and abundant, unlike platinum. Nitrogen-and boron-doped carbon nanoparticles have shown promising results for both OER and HER (Yeh et al., 2018). Ndoped CNTs were used to load palladium which produced comparable results to commercial-grade Pt/C (Sarkar and Peter, 2018). While reducing the costs promoted research in cathode development, the anode faced a different problem.

The electrode material and the corresponding catalyst are constantly subject to high anodic potentials of more than 1.5 V, which leads to material degradation. This degradation is much more significant at potentials greater than 1.8 V (Reier et al., 2014). Recent developments in OER electrocatalysts include the reduction of noble-metal content by producing alloys with transition metal oxides with the primary catalysts (IrO2 and RuO2), such as oxides of Ti, Sn, Ta, Nb, Sb, Pb, and Mn. Increased stability in electrode material was observed with a mixture of 20% Nb2O5 with RuO2, called the Adams method (Puthiyapura et al., 2013). Other studies showed a strengthened OER electrode with a mixture of nanostructured SnO2-IrO2-TaO5 (Ardizzone et al., 2006). Many other combinations of metal oxides studied toward improving OER and reducing degradation of anode electrocatalysts are mentioned in Shiva Kumar and Himabindu (2019). Apart from the electrode and electrocatalyst, the corrosive conditions of a PEM pose an additional problem of membrane degradation. One of the primary ways to control the membrane is by providing even compression over the membrane surface due to mechanical stresses (Khatib et al., 2019). Reinforcing the membranes with electrically nonconducting material has also shown a positive impact on the membrane stability and lifetime (Liu et al., 2015a). Although thin membrane layers have been suggested to decrease ohmic losses and to obtain higher current densities, they are unsuitable for highly intensive and large-scale PEM applications (Xu et al., 2012). Novel methods such as operating the PEM cells in high recycling rates (>90%) have shown high performance efficiency with minimal losses. These novel methods have also shown increased membrane material recovery, reducing the environmental impact of the PEM process (Carmo et al., 2019).

8.2.3 Solid oxide electrolyzers


Solid oxide electrolyzers (SOEs) are the most recent development in electrolyzer technology that work on the principle of anion exchange in the form of O2− (Fig. 8.3). The concept of SOE was developed in the 1980s with the aim of achieving 100% electrical efficiency and pure gaseous products with low operating potentials of approximately 1.07 V. The development of SOE is in the very early stages and is a long way from commercialization and large-scale operation. SOE is operated under high-pressure and high-temperature (500°C– 850°C) conditions in tubular electrolyzers that contribute to the higher efficiency. Although the higher temperatures demand high energy for the operation, the corresponding electrical energy required for water splitting is reduced significantly (as reflected in the low potential requirement) and thereby decreases the overall energy demand of the process. Therefore a renewable or recycled heat source for SOE makes it a highly attractive and adaptive method for hydrogen production. The operating principle of SOE shows that steam and recycled H2 feed at the cathode to reduce water and produce hydrogen as shown in Eq. (8.11). The oxide anions formed at the cathode go through the solid electrolyte to the anode, where they recombine, according to Eq. (8.10), to form oxygen (Ursúa et al., 2012).

Figure 8.3 Schematic representation of solid oxide electrolyzers (Khatib et al., 2019).

Electrolysis reactions include the following: Anode reaction

(8.10)

Cathode reaction

(8.11)


The SOE demands porous electrodes to maximize the between the electrode surface and the gaseous reactant and product species. The anode is made of yttria-stabilized zirconia (YSZ) composited with various perovskites as lanthanum manganites (LaMnO3), ferrites (LaFeO3), or cobaltites (LaCoO3) in part substituted with strontium. The cathode is also composite cermet made of nickel and YSZ. The SOE consists of a solid ceramic electrolyte made of YSZ, which is known for its high stability and ionic conductivity at the high temperatures of SOE cells. A total of 17.6 N L/h of hydrogen has been produced with this cell configuration at 1000°C, 0.4 A/cm², and 39.3 W of applied power.

A stack consisting of 1000 electrolysis cells produced up to 600 N L/h H2-gas (Yan and Hino, 2016). Attempts by other researchers to improve hydrogen production rates include SOE cells planar configuration and the use of calciastabilized zirconia (Zahid et al., 2010). To take advantage of reduced energy demand through recycled or renewable heat supply for the operation of SOE, many reactor configurations have suggested integrating the SOE cell stacks with solar or nuclear power sources. The waste heat from nuclear power production facilities is best utilized for the generation of another energy source such as hydrogen. Solar heat is used for evaporation and superheating of water with the help of a solar simulator and steam generator that are able to generate electrical cell stack efficiencies of about 93% (Schiller et al., 2019). Other areas where the SOE-based research is focused on are (1) cheaper alternatives for the separation of hydrogen and steam after they are released from the cathode chamber and (2) electrolysis cells steadiness that is addressed via reduced electrolyte degradation and electrode deactivation (Zahid et al., 2010).


Most recent developments in the SOE in of power-to-gas have focused on reactor configuration and strategizing the integration with methanation reactors. Some studies have carried out technoeconomic analyses in of electricity price fluctuations and long-term operation (Zhang et al., 2020). A study by Wang et al. was published in 2019 describing a stack arrangement operation for 6000-h under laboratory isothermal conditions. The aim of the study was to understand the increase in stack temperature due to degradation, which ranked power-tomethane second to power-to-hydrogen. However, the study concluded that solid oxide electrolysis will be most suitable for PtG in of long-term operation under a coelectrolysis system (Wang et al., 2019a). Other system-level research has shown that integrating SOE with oxyfuel combustion (i.e., using pure oxygen instead of air to burn carbonaceous materials, a near CO2-free system; Mathieu, 2010) allows for high exergy efficiencies of about 65%–75% (Eveloy, 2019). However, this technology needs to be traded off with low electricity

prices and that requires a systemic change in how electricity is sourced (Wang et al., 2019b).

8.2.4 Fixed-bed methanation reactors


Fixed-bed reactors are the main type of catalytic reactors for large-scale chemical synthesis. The process is characterized by different gaseous chemical species reacting on the catalyst surface placed in a fixed position inside the reactor. There are two main types of fixed-bed reactors (1) adiabatic and (2) multitubular fixed-bed reactors (Fig. 8.4). Adiabatic reactors are used where there is only one main reaction pathway, whereas the multitubular fixed-bed reactors are utilized for product specificity and separation. The reactant gases flow uniformly over the fixed bed in an adiabatic process, whereas in the multitubular reactor the feed gas follows the path of a heat carrier that allows for control of different temperature profiles around the fixed tubular catalysts (Eigenberger, 1992). The methanation reactions described in Eqs. (8.3) and (8.4) are evidently highly exothermic and therefore require better control over temperature. Several configurations of the two types of fixed-bed reactors have been explored over the last two to three decades. Many of the methanation reactors discussed in research articles deal with the methanation of syngas as a feed gas, however, a similar concept was applied to PtG where CO2, CO, and H2 come from two different sources instead of a gasification process. More recently, plants have started to work with methanation for biogas upgradation from wastewater treatment plants, bioethanol production plants, and some use bottled gas as a source as CO2 (Thema et al., 2019).

Figure 8.4 Schematic representation of the two types of fixed-bed reactors: (A) adiabatic fixed-bed reactor; (B) multitubular fixed-bed reactor.


Considering that large quantities of heat are released in the methanation process, isothermal fixed-bed reactors were first tried in a series of at least two isothermal fixed-bed reactors. Maximum reaction control and heat transfer were achieved with a combination of outlet gas recycling and intermediate cooling steps (Kopyscinski et al., 2010). Operating temperature and pressure for the fixed-bed methanation reactors in series vary in the ranges of 200°C–800°C and 25–80 bar. Other configurations include (1) partial product recycling, (2) “one through” feed gas flow (GF) without preheating, and (3) series of isothermal and adiabatic reactors, where one acts as a heat exchanger and the other increases the product yield (Schaaf et al., 2014). In 2014, Schaaf and group designed a lab-scale fixedbed reactor for the methanation of CO and CO2 and followed it with large-scale plant designs. The pilot reactor was operated under various conditions of gas hourly space velocity, reactor inlet gas temperature, and reaction temperature. Target uniform temperatures of 400°C, 450°C, and 500°C were obtained on all of the catalysts tested placed between two inert material layers. The final outlet gas composition was CH4: 82%; H2: 13%; CO2: 4%; CO: 0.02%; and ethane: 0.01%. Two plant design concepts were described based on configurations (2) and (3) described earlier (Schaaf et al., 2014).


Fixed-bed reactor design concepts have long been explored and the state of art has been published in many scientific research articles, review articles, and

books. In recent times, the scientific community working toward chemical methanation has focused its research interests toward reactor operation optimization, such as start up control and dynamic operation that is more practical and economical to take up PtG. This is important because RE such as solar and wind power are not consistent in electrical power production and require the subsequent processes to be dynamic and adaptive to the availability and production of renewable electricity. Fixed-bed methanation reactors for PtG must be equipped with appropriate catalysts that can be reactivated after being exposed to long hours of nonreacting carbonaceous reactants, that is, CO2 or CO. The reactor design and materials must frequent starting and stopping of methanation without start-up delays and inefficiencies (Rönsch et al., 2017). Bremer et al. (2017) modeled a two-dimensional fixed-bed tube reactor for CO2 methanation. The model was sustained on the underlying exothermic Sabatier reaction scheme. The authors were able to control dynamic hot spot formation inside the catalyst bed, while proving the feasibility of dynamic carbon dioxide methanation. Kreitz et al. (2019) applied periodic concentration forcing of CO2 in fixed-bed methanation reactors set up with Ni/Al2O3 catalyst. The results indicated no methane formation during the CO2 concentration forcing phase, whereas fast hydrogenation reaction rates were observed upon switching to the methanation phase after turning on the H2 supply (Kreitz et al., 2019). Several other models have been developed for dynamic operation of CO and CO2 methanation in fixed-bed reactors that are yet to find a laboratory application followed by pilot and large-scale studies.

8.2.5 Fluidized bed methanation reactors


Fluidized bed reactors are a popular method for carrying out highly exothermic reactions both in lab and large scale, such as gasification of biomass,

methanation of CO2, etc. The fluidized bed reactor was primarily designed to overcome the heat transfer efficiency and temperature control issues faced in fixed-bed reactors (Lappas and Heracleous, 2011). The fluidized bed reactor is based on the principle of turbulent GF and rapid circulation of heat through inert heat carriers and catalysts. This reduces the exchange area and the costs for gas compression, which helps in bringing down the overall cost of construction and operation. There are three main types of fluidized bed reactors that are differentiated based on the relative flow rates of catalyst setting (CS) and GF (1) bubbling reactor (CS=GF), (2) circulating (CS>GF), and (3) transport (CS

Early studies on fluidized bed reactors showed methanation of CO and CO2 to enrich biomass-based syngas to synthetic natural gas and are available as research and review articles (Calbry-Muzyka et al., 2019; Sun et al., 2018). The state-of-the-art was previously established in the form of the COMFLUX process developed in Austria in 2009 (Götz et al., 2014). More recent work in methanation deals with methanation of CO2-rich polluted gas/biogas with little or no CO in the feed gas without prior separation of CO2. Biogas upgrading using thermochemical methanation was demonstrated in pilot scale for the first time using a fixed-bed methanation reactor and SOE cell in Denmark (Hansen et al., 2016). Recently, a bubbling fluidized bed reactor named COSYMA (Container-based SYstem for MethAnation) (Fig. 8.5) showed stable operations of over 1000 h with biogas feed and 96% methane conversion efficiency with the final product containing about 88% CH4, 11% H2, and 1% CO2 (Witte et al., 2019).

Figure 8.5 Simplified flowsheet of the Container-based SYstem for MethAnation set-up (Witte et al., 2019).

The Ni-based catalysts used in the chemical CO2 methanation process face deactivation issues due to carbon deposition on the catalyst surface. The carbon particles block the catalytic pores and deteriorate catalytic activity, which increased considerably with high-temperature operations (Fei et al., 2018). In a recent study, researchers used an innovative approach to improve Ni catalyst efficiency by over 100% using SiO2 spheres and SiO2 spheres coated with TiO2 as materials for Ni and NiCo bimetal catalysts (Jia et al., 2019). Similar results were exhibited in another study where TiO2 was used as an additive material for Ni-Al2O3 catalyst. TiO2 worked as a physical roadblock to hydrocarbon adsorption and thus decomposition on the catalyst surface, and the oxygen active sites on Ni-Ti/Al2O3 promoted the removal of carbon from the Ni particles (Feng et al., 2019).


The bubbling-type reactor has taken the center stage in of the development of the fluidized bed methanation reactor research. In bubbling-type fluidization, the reactant gas es through a bed of catalyst particles, creating two phases, namely emulsion (dense) and bubbles (voids). The bubbles are of low solid density and on an average move upward faster than the gas and increase the catalyst particles. At the same time, bubbles rise and move downward in the emulsion zones. This results in a circulation pattern which enhances mixing of different flows and provides efficient heat transfer from the main reaction zone to both the inlet and outlet. As catalysts, gases, and bubbles are all in constant movement, the formation of laminar gas films is disrupted and leads to a highly efficient heat transfer mechanism among the reactant and catalyst surfaces. However, this mechanism still fails to observe a perfect isothermal system owing to the rise in temperature at the inlet due to the intense reaction (Seemann and Thunman, 2019).

More recent studies are now focusing on simulation of bubbling fluidized bed reactors using computational fluid dynamics in combination with the discrete element method (Li et al., 2019). Comparisons of these studies are done with two-fluid models and kinetic theory (Li et al., 2020). Whatever the method and approach for the reactor design may be, the aim of these researchers has been to create a perfect isothermal reactor by studying the heat transfer mechanisms and removing the heat as soon as the reaction generates it (Nam et al., 2020).

8.2.6 Three-phase reactor


A three-phase reactor, otherwise called a slurry bubble column reactor (SBCR), uses a mixture of gas, liquid, and solid during its operation. The gas phase may consist of both feed and product, the solid phase is the catalyst material, while the liquid feed blended with catalyst forms a slurry, to achieve efficient heat transfer, making the liquid phase. This is mainly due to the high heat capacity of the liquid and effective mixing of the slurry phase that makes SBCRs compatible with highly exothermic reactions such as the methanation process. There are typically three methods of operating an SBCR based on GF regimes: (1) homogeneous: low GF velocity with small bubbles; (2) heterogeneous: increased GF velocity such that some of the bubbles coalesce; and (3) slug flow: high GF velocities with bubbles as large as the diameter of the reactor. A heterogeneous flow regime is preferred for a highly exothermic reaction such as CO2 methanation due to higher back mixing, gas holdup, and mass transfer (Lefebvre, 2019). The main motivation behind the development of three-phase methanation (3PM) is its ability to maintain isothermal conditions over long periods of operational and down time, allowing the methanation process to be flexible and dynamic, making it suitable for PtG applications. Some of the challenges with three-phase methanation reactors include gas–liquid mass transfer limitations and also decomposition and evaporation of the heat transfer liquid (Götz et al., 2016).


The research on 3PM reactors is in its very early stage, where a few research groups around the world are working on using bench-scale setups aiming to scale it up. Many of the studies use an autoclave as the main reactor body where different gases are fed through a feed tank and the product gases are collected in a condensate tank. Dibenzyltoluene is one of the popular liquids applied for 3PM reactions due to its wide range of gas solubilities and low vapor pressure at reaction temperatures of around 250°C and 320°C (Fig. 8.6). Other liquids that are commonly used and evaluated are squalane, octadecane, polydimethylsiloxane, and ionic liquids, namely {butyl-trimethyl-ammonium bis(trifluoromethylsulfonyl) imide [N1114] [BTA], 1-methyl-1- propylpiperidinium bis(trifluoromethylsulfonyl) imide [PMPip] [BTA], and 1-ethyl-3methylimidazolium trifluoromethanesulfonate [EMIM] [Tf]} that showed positive results on a 3PM operation (Lefebvre et al., 2018; Götz et al., 2013).

Figure 8.6 Schematic diagram of the three-phase reactor (Götz et al., 2014).

8.2.7 Micro(channel) reactors

Reactors that are constructed at submillimeter range dimensions and are used for chemical production applications would be considered as microreactors. These reactors use microchannel flow regimes and take advantage of the high mass and heat transfer efficiencies that occur at this scale due to the high area-to-volume ratios. Microchannel reactors can be classified as a polytropic process, where temperature and pressure are maintained constant throughout the process. The microchannel reactor configuration s the plug flow pattern, which helps to obtain the ideal reactor performance for the methanation reaction. Due to the a high precision control of heat transfer, operational costs such as catalyst deactivation/replacement can be avoided. Such reactors require a high level of scale-up procedures such as running numerous microreactors of the same size in parallel, unlike other reactors that are only increased in their dimensions. The conventional method of scaling up does not allow for recreating similar process efficiencies that are achieved in lab or pilot-scale reactors. This could be mainly due the pressure drop that occurs within the huge reactors that is not observed in smaller scale operations. The typical operating temperature for the microchannel design is between 250°C–500°C. The catalysts are coated on to the walls of the reactor and are typically a combination of alternative metal catalysts (i.e., Ni, Ru, Rh, and Co) with various oxide materials (i.e., TiO2, SiO2, MgO, and Al2O3). The combination of Ru-TiO2 is one of the most popular catalysts employed for microreactors. Brooks and collaborators have experimented by coating the powdered form of this catalyst on to FeCrAlY metal felt material instead directly on to the walls of the channels (Fig. 8.7). The porous metal felt coated with the catalyst is placed flat on the base of the channel on which the feed gases were introduced. The reactor design with two sets of microchannels placed beside each other with the coolant channels at either side of the microchannels allowed for efficient heat exchanging system (Brooks et al., 2007).

Figure 8.7 Engineering drawing of the Sabatier microchannel reactor assembly (Brooks et al., 2007).

In another study, working with microchannel reactors revealed the pathway of the methanation of CO to CH4 is by first oxidizing CO to CO2 at lower temperatures with oxygen followed by high-temperature CO2 reduction to CH4 in the presence of hydrogen. The research group observed complete removal and reduction of CO to CH4 with Ru/SiO2 combination catalyst coated directly on to the stainless steel microchannel reactors (Görke et al., 2005). Syngas methanation has been proven with high conversion and selectivity over Ni catalyst in a microchannel reactor in the order of a time of milliseconds (Liu et al., 2012). The authors created novel metal–ceramics complex substrate that was employed as the catalyst , on which the Ni catalyst was impregnated. The substrate formed a stable and active catalyst coating, which was considered reliable for engineering applications of microchannel reactors. Isothermal temperatures of about 550°C with consistent CO conversion efficiencies of about 98% were achieved. Microchannel reactors are a promising technology that paved the way to creative microstructured fixed and fluidized bed reactors. A microstructured fixed-bed reactor contains packed catalyst modules in milli/microscale cavities separated by similar-sized catalysts containing coolant material (Gruber et al., 2018). However, a recent study exhibited the dynamic simulation of a microstructured fixed-bed reactor for methanation which revealed catalyst deactivation to be a problem even at the microscale and concluded that a multistage reactor design needs to be studied for the purpose (Kreitz et al., 2019b). Another recent study investigated a biogas upgrading plant using a two-stage microstructured heat exchanger reactor (gas preheating, catalytic reaction, and water condensation), which showed promising results for gas grid integration (CH4 >95%, H2 <5%, and CO2 <2.5%) (Guilera et al., 2020).

8.3 Bioelectrochemical methanation

Bioelectrochemical systems (BES), such as microbial fuel cells (MFCs), were one of the first to be studied in the 1980s and 1990s for wastewater treatment systems that produced electricity (Liu et al., 2004). However, the technology could not find large-scale applications due to its low production capacities. Carbon capture or carbon recycling was prioritized over electricity production with the start of the new millennium. The MFCs were modified to microbial electrolysis cell (MEC) and microbial electrosynthesis system (MES), where electricity was input to the bioelectrochemical cell in order to carry out the nonspontaneous reactions such as hydrogen evolution and CO2 reduction, respectively (Fig. 8.8) (Guo et al., 2013; Januszewska et al., 2014; van EertenJansen et al., 2015; Bajracharya et al., 2016; Cai et al., 2016; Zhao et al., 2016; Aryal et al., 2017b; Bajracharya et al., 2017). The microbial catalysts and the corresponding potential determine the end product of the bioelectrochemical reaction. Hydrogen production occurs at a cathode potential of −0.414 V, whereas the microorganisms, as catalysts, generate a cathode potential of about −0.300 V by feeding on organics in the bioelectrochemical cell. It is possible to form H2-gas with only a little additional potential through the electrolysis process aided by microorganisms (Liu et al., 2005). However, it was observed that methane was an unavoidable byproduct of electrolytic hydrogen production due to the presence of electro-active methanogens on the cathode. It was realized that methane as a fuel is more practical than hydrogen as it is cheaper and easier to produce, store, transport, and use. For example, in comparison to methane (36 MJ/m³) hydrogen gas energy density is very low (11 MJ/m³), giving some disadvantages in its use as a transport fuel (Balat et al., 2008). Methane production via carbon dioxide reduction was later demonstrated by the same research group in 2008 (Call and Logan, 2008).

Figure 8.8 A schematic representation of a membrane-less microbial electrosynthesis system and microbial electrolysis cell for the treatment of wastewater and CO2 (Nelabhotla and Dinamarca, 2018).

The two main pathways identified for bioelectrochemical methane production are indirect and direct electron transfer. As the names suggest, indirect electron transfer (IET) uses hydrogen as an intermediary chemical species to transfer electron [Eqs. (8.12)– (8.14)] whereas in the latter, electrons are transferred directly from the electrode surface to the microorganisms that produce methane [Eqs. (8.15) and (8.16)]. With direct electron transfer (DET), methane generation can be achieved in PtG systems without a separate hydrogen-producing electrolysis unit. Generally, the electron transfer pathway, the electrode material, and the electrocatalysts determines the electrochemical cell efficiency. Indirect electron transfer:

(8.12)

(8.13)

(8.14)

Direct electron transfer:

(8.15)

(8.16)

Apart from direct and IET for methane production in the bioelectrochemical cell, methane can also be produced by heterotrophic and hydrogenotrophic bacteria that either are attached to the cathode surface or present in the bulk media. All the possible methane production pathways are summarized in Fig. 8.9.

Figure 8.9 Possible mechanisms of methane production in integrated MEC/S-AD systems: methane production in the bulk by heterotrophic bacteria otherwise called exoelectrogenic fermentation bacteria; indirect electron transfer via hydrogenotrophic methanogenic archaea that are in bulk as well as biofilm; and direct electron transfer by electrotrophic methanogenic archaea ( Feng et al., 2018).

8.3.1 Direct electron transfer

The transfer of electrons without any intermediary species between the electrode material and the biofilm attached on the surface is considered as DET (Siegert et al., 2014). Due to the absence of intermediary species, the diffusional limitations are reduced and, due to the biofilm , electrode overpotentials are reduced. These factors help to improve the process efficiency and product yield in bioelectrochemical methane production. Some studies have observed complete stoppage of hydrogen due to the activity of a mixed population of microorganisms present on the biofilm (Mueller, 2012; Xu et al., 2014). This paradigm shift in the mechanisms of electron transfer significantly impacts the “modeling and design of anaerobic wastewater reactors and the understanding of how methanogenic communities respond to environmental perturbations” (Morita et al., 2011). Many studies have been conducted to understand the mechanism, biofilm structure, and behavior of IET, with hydrogen and formate acting as electron carriers between cells in close syntrophic association (Thiele et al., 1988). However, earlier studies revealed the presence of conductive pili that are of nanoscale and form a network of connections among the electroactive microorganisms to transfer electrons directly (DET). Other studies showed the presence of flagellum-like appendages to promote electron transfer as well as other syntrophic processes such as energy exchange. Control studies of MES that were operated with and without the presence of biofilm on the electrode surface were used to DET (Gorby et al., 2006). These studies showed higher current densities in the presence of biofilm and showed methane as the primary

product when fed with CO2 (Cheng et al., 2009). In 2014, Rotaru and group demonstrated electron conductivity using granular active carbon with a combination of pili-deficient Geobacter metallireducens and methanogenic bacteria (Rotaru et al., 2014). A correlation factor (R²)=0.67 was obtain in relation to the enhanced granule conductivity through DET for methane production and abundance of Geobacter species (Shrestha et al., 2014). Another control study of an MEC-AD integrated reactor with and without Geobacter species revealed lower methane yield in the system without Geobacter. This concludes that electron transfer between methanogens and Geobacter is an important reaction that is significantly improved through DET for CO2 reduction to methane (Yin et al., 2016). Research on electron transfer mechanisms generated significant interest in MES as it was then possible to carry out efficient electrochemical methane production with cheaper electrode materials such as carbon cloth (Chen et al., 2014a), biochar (Chen et al., 2014b), and magnetite (Liu et al., 2015b). The intricate and highly available surface area of these electrode materials allows electron conductivity between the electrode material and microorganisms, as well as among the microorganisms in the biofilm. Such a biofilm network inside and on the surface of a carbon electrode can save energy by avoiding pili generation (Zhao et al., 2015). The carbon materials employed for improving conductivity have also proven to protect the biofilm network from destabilizing environments such as acid impacts and high hydrogen partial pressures if and when they occur (Zhao et al., 2017).

8.3.2 Biocathodes

Biocathodes in the case of MES are mainly based on biofilm cultivation over the cathode surface using a mixed or pure culture of bacteria to carry out the desired reaction at a particular cathode potential. They served as cheaper and sometimes even better alternatives to expensive noble metal catalysts such as platinum. Bioelectrochemical hydrogen production was demonstrated for the first time by using a bioanode developed in an MFC as a biocathode in an MEC. This was done through inversion of polarity and slow adaptation of the microorganisms on

the electrodes from oxidizing to reducing environments (Rozendal et al., 2008). Later, in 2010, a graphite block biocathode was developed for MES systems demonstrating CO2 reduction for chemical production (Nevin et al., 2010). This was followed by several innovative methods of biocathode development serving toward electrosynthesis of various chemicals and biofuels (Aryal et al., 2017a). The method in which biofilm is developed on the electrode surface determines the biofilm composition, their symbiotic relation, electron transfer efficiency, and the product yield. Biofilms are strengthened by different interactive forces such as H-bonding, electrostatic attraction, and van der Waals interaction (Jourdin et al., 2016). However, it is considered that a mixed and diverse microbial culture enriched within the MES reactor, through a seed inoculum or the feed itself, is highly stable and robust toward different physiological stresses in the reactor as compared to pure cultures (Ganigue et al., 2015). However, there is no conclusive evidence presented as to how electrons are transferred to the biofilm network based on the conductive pili. Some researchers have suggested metal-like conductivity (Malvankar et al., 2011), whereas others have suggested transfer of electron through c-type cytochromes present on the surface of microorganisms (Lovley et al., 2011). Various cathode materials have been explored for MES, such as graphite, carbon cloth, carbon plates, etc., which are fundamentally 2D structures that limit the potential of biofilm cover/spread and activity. Carbon felt is a sponge-like intricate woven structure that allows for a highly interactive environment for the biofilm with the electrode material and other microorganisms as well. It also allows for varied microbial species, depending on their location of attachment on the electrode surface that determines its exposure to different physico-chemical conditions within the reactor environment. Carbon felt was used in MES for the first time by Jiang and group in 2013, where MES product diversity was conclusively presented based on different cathode potentials. The study showed methane and hydrogen were produced at a cathode potential range of −0.65 to −0.75 V (vs. SHE), methane, hydrogen, and acetic acid at −0.75 V (vs. SHE), and methane and acetic acid at −0.95 V (vs. SHE) (Jiang et al., 2013). More recently another research group observed thick biofilm covering over the carbon felt electrode generating high current densities through biofilm adaptation (Jourdin et al., 2018). Similar results were obtained by Nelabhotla et al. (2019) through biofilm adaptation for electrochemical methane production (Nelabhotla et al., 2019).

8.3.3 Reactor configurations

MES is a system derived from MFCs, MECs, and electrolysis that is highly reliant on two different electrolytes in the two chambers (anode and cathode). The electrolysis process also releases two different gases at the two electrode terminals and requires separation to maintain gas purity. Therefore these reactors need selective permeable membranes such as a proton/cation exchange membrane. These membranes, although benefiting the electrochemical process, have several disadvantages such as reduced mass transfer efficiency, current densities, and fouling. Due to the prolonged usage of such membranes in electrochemical systems, the earlier MES were also developed as two-chamber reactors. However, it was realized through several studies that MES are highly efficient when operated as single-chamber reactors due to ohmic resistance, minimized pH gradient, and higher current density (Guo et al., 2017). An indepth review of all the single-chambered microbial electrolysis/electrosynthesis systems has been presented by Nelabhotla and Dinamarca (2018). The important factors that contribute to MES reactor design/configuration are its feed stream and the desired product. MES used for biochemical synthesis needs to be designed differently to MES for biofuel synthesis. For example, biochemical synthesis would require pure CO2 that could be sourced as carbon capture canisters and be converted to the desired chemical. On the other hand, MES for biofuel/methane production is shown to be hugely beneficial when used for biogas upgrading where the feed stream, that is, biogas, contains 30%–50% CO2 (Xu et al., 2014). Several studies were published demonstrating the process parameters and reactor designs for integrated AD-MEC/MES systems (Fig. 8.10) (Nelabhotla and Dinamarca, 2019; Moreno et al., 2016; Bo et al., 2014). It is important to identify the source of electrons for CO2 reduction in these integrated systems. In an MEC, electrons are obtained via organic/acetate degradation, which contributes additional CO2 production in the anodic chamber that is reduced to methane in the cathodic chamber. In an MES, the electrons are contributed via partial water splitting reducing the CO2 that is part of the feed biogas stream. Although both (MEC and MES) integrated systems contribute to increased methane yield, a clear distinction must be made with regards to carbon

dioxide reduction both chemically as well as volumetrically. While increasing the methane production rate, the AD integration with MEC contributes to additional acetate/organic consumption (Zhao et al., 2014), whereas AD integration with MES contributes lower CO2 emissions. Recent publications have suggested anodic degradation of COD as an electron donor in an MES in addition to CO2 reduction (Nelabhotla et al., 2019); however, much research still needs to be carried out to scale-up such integrated reactors.

Figure 8.10 Bioelectrochemical systems integrated with AD as a posttreatment technology ( Li et al., 2015).

8.4 Challenges and future prospects

The increasing share of RE sources needs flexible forms of capacity storage due to the fluctuating input. PtG for methanation is therefore a good fit to contribute to the gradual transformation from a fossil to postfossil society. Methane, a green energy molecule, is used for heat, electricity generation, and transport, but it is also a feedstock for several industrial process, and its production also injects CO2 into the renewable cycle loop, thus reducing greenhouse gas emissions. These reasons and the necessity to strength energy security and diversify supply give PtG-CH4 good prospects for further development and full implementation. Nevertheless, the proven concept technology as outlined in this chapter is limited by low efficiencies and high cost. Key technological aspects for both hydrogen production and CO2 methanation are under continuous development, and a steep learning process has been taking place through many projects across Europe in recent years. These upscaling projects stress the importance of suppleness, high efficiency, and low CAPEX that at the industrial level are in the end the most important criteria for succeeding. It is perhaps therefore that at the integrated value chain concept is where PtG becomes long-term economically feasible and self-sustaining, especially where industrialized parks exist and a symbiosis between companies is possible. Heat, metals, waste sludge, CO2, CO, H2, and methane can be raw materials and products available at short distances. The use of existing infrastructure and distribution systems will further ensure this vision. That may sound idealized, or perhaps not, given the more mature environmental awareness that is moving companies to seek partnerships, as such PtG methanation, not just for the sake of energy conversion, but if conceived properly, could be at the heart of waste management, nutrient and metals recovery, energy for transport, and innovation hubs. Methanation requires both the supply of electron equivalents and an inorganic carbon source. This in an appropriate stoichiometric ratio. CO2 supply can be obtained elsewhere, with an increasing CO2 concentration in the atmosphere and consequently growing awareness of environmental challenges, industries with their exhaust gases are more than welling to get rid of this greenhouse gas.

It is, however, more challenging to obtain the available equivalent electrons, which in electrocatalytic systems are solely derived from electrochemical water splitting at a minimum theoretical voltage of 1.2 V, to produce hydrogen gas as an indirect electron carrier to methane. Alternative biological mediated systems can use diverse electron equivalent sources as water, organic substances, ammonium, and sulfides. BESs are in the earlier stage of development, but they have the advantage of using biocatalyst, living microorganisms that drive chemical reactions under considerably lower potentials. Another important characteristic of BES is that electrons (or equivalents) can be directed toward the generation of methane in a single reactor. There is, however, much to advance in this field, and most studies so far have focused on the biochemical reactions that take place at the cathode, while the fundamental studies in bioanodes are insufficient. Another important advantage of biological-mediated methanation is the higher tolerance to impurities, which facilitates the use of biomass such as wood, waste sludge, and general organic waste. Thus there is a connection to an important source of electrons. Gasification or pyrolysis of such organic biomass resources can provide equivalent electrons in the form of H2 and CO gases. It is therefore an important link between PtG for methanation and syngas fermentation, which is of special relevance in existing biogas plants due to the increasing restrictions on disposing of digested sludge (e.g., as in an agricultural field by reason of microplastics and chemicals getting into our food production). Often PtG methanation has been used to signify water electrolysis followed by hydrogenotrophic methanogenesis, which also is the main theme of this chapter, and equivalents from water are used to convert CO2 into methane. In biogas plants methane is most often upgraded in two steps to remove CO2 and H2S; plants are also limited by the ammonium concentration, which is toxic for the bacterial culture at high pH and needs to be handled downstream (e.g., ammonia stripping). BES could potentially solve at least partially these challenges by oxidizing sulfides and ammonium, and reducing CO2 to methane in a single operational unit. The need for energy storage capacity, energy source flexibility, and security gives renewable gas undoubtedly an important role in the future European energy system. Process technologies to realize power-to-gas for methanation are the most adequate to meet these demands, given the already existing infrastructure, distribution systems, and in some part policies (e.g., as cooperation agreements between food producers and biogas plant owners) as found in biogas plants.

References

Antenucci and Sansavini, 2019 Antenucci A, Sansavini G. Extensive CO2 recycling in power systems via power-to-gas and network storage. Renew Sustain Energy Rev. 2019;100:33–43 https://doi.org/10.1016/j.rser.2018.10.020. Anzola-Rojas et al., 2018 Anzola-Rojas M, Mateos R, Sotres A, et al. Microbial electrosynthesis (MES) from CO2 is resilient to fluctuations in renewable energy supply. Energy Convers Manage. 2018;177:272–279 https://doi.org/10.1016/j.enconman.2018.09.064. Appl, 1999 Appl M. Ammonia: Principles and Industrial Practice Chichester: Weinheim; 1999. Ardizzone et al., 2006 Ardizzone S, Bianchi CL, Cappelletti G, et al. Composite ternary SnO2–IrO2–Ta2O5 oxide electrocatalysts. J Electroanal Chem. 2006;589:160–166 Available from: https://doi.org/10.1016/j.jelechem.2006.02.004. Aryal et al., 2017a Aryal N, Ammam F, Patil SA, Pant D. An overview of cathode materials for microbial electrosynthesis of chemicals from carbon dioxide. Green Chem. 2017a;19:5748–5760 https://doi.org/10.1039/C7GC01801K. Aryal et al., 2017b Aryal, N., Zhang, T., Tremblay, P.-L., 2017b. Microbial Electrosynthesis for Acetate Production From Carbon Dioxide: Innovative Biocatalysts Leading to Enhanced Performance. Bajracharya et al., 2016 Bajracharya S, Sharma M, Mohanakrishna G, et al. An overview on emerging bioelectrochemical systems (BESs): technology for sustainable electricity, waste remediation, resource recovery, chemical production and beyond. Renew Energy. 2016;98:153–170 https://doi.org/10.1016/j.renene.2016.03.002. Bajracharya et al., 2017 Bajracharya S, Van Den Burg B, Vanbroekhoven K, et

al. Electrochimica acta In situ acetate separation in microbial electrosynthesis from CO2 using ion-exchange resin. Electrochim Acta. 2017;237:267–275 https://doi.org/10.1016/j.electacta.2017.03.209. Balat et al., 2008 Balat M, Balat H, Oz C. Progress in bioethanol processing. Prog Energy Combust Sci. 2008;34:551–573 https://doi.org/10.1016/j.pecs.2007.11.001. Barbir, 2005 Barbir F. PEM electrolysis for production of hydrogen from renewable energy sources. Sol Energy. 2005;78:661–669 https://doi.org/10.1016/j.solener.2004.09.003. Biesemans, 2016 Biesemans, M., 2016. Biogas Upgrading by Means of Simultaneous Electrochemical CO2 Removal and Biomethanation. Bo et al., 2014 Bo T, Zhu X, Zhang L, et al. A new upgraded biogas production process: coupling microbial electrolysis cell and anaerobic digestion in singlechamber, barrel-shape stainless steel reactor. Electrochem Commun. 2014;45:67–70 https://doi.org/10.1016/j.elecom.2014.05.026. Braathen et al., 2019 Braathen, M., Sømme, H.O., Kaurin, M., Helland, A., Jahren, T.Ø., 2019. Northern Lights: Konsekvensvurdering Med Hensyn På Fiskeri Og Marint Biologisk Mangfold Vest For Grunnlinjen. Bremer et al., 2017 Bremer J, Rätze K, Sundmacher K. CO2 Methanation: optimal start-up control of a fixed bed reactor for power-to-gas applications. AlChe Journal. 2017;63:23–31. Breyer et al., 2015 Breyer C, Tsupari E, Tikka V, Vainikka P. Power-to-gas as an emerging profitable business through creating an integrated value chain. Energy Proc. 2015;73:182–189 https://doi.org/10.1016/j.egypro.2015.07.668. Brooks et al., 2007 Brooks KP, Hu J, Zhu H, Kee RJ. Methanation of carbon dioxide by hydrogen reduction using the Sabatier process in microchannel reactors. Chem Eng Sci. 2007;62:1161–1170 https://doi.org/10.1016/j.ces.2006.11.020. Buckley, 1974 Buckley D. The oxygen electrode Part 4 Lowering of the overvoltage for oxygen evolution at noble metal electrodes in the presence of ruthenium salts. J Electroanal Chem. 1974;52:433–442

https://doi.org/10.1016/0368-1874(74)85058-6. Cai et al., 2016 Cai W, Liu W, Yang C, et al. Biocathodic methanogenic community in an integrated anaerobic digestion and microbial electrolysis system for enhancement of methane production from waste sludge. ACS Sustain Chem Eng. 2016;4:6b01221 https://doi.org/10.1021/acssuschemeng.6b01221 acssuschemeng. Calbry-Muzyka et al., 2019 Calbry-Muzyka AS, Gantenbein A, Schneebeli J, et al. Deep removal of sulfur and trace organic compounds from biogas to protect a catalytic methanation reactor. Chem Eng J. 2019;360:577–590 https://doi.org/10.1016/j.cej.2018.12.012. Call and Logan, 2008 Call DF, Logan BE. Hydrogen production in a single chamber microbial electrolysis cell lacking a membrane. Environ Sci Technol. 2008;42:3401–3406 https://doi.org/10.1021/es8001822. Carmo et al., 2013 Carmo M, Fritz DL, Mergel J, Stolten D. A comprehensive review on PEM water electrolysis. Int J Hydrog Energy. 2013;38:4901–4934 https://doi.org/10.1016/j.ijhydene.2013.01.151. Carmo et al., 2019 Carmo M, Keeley GP, Holtz D, et al. PEM water electrolysis: innovative approaches towards catalyst separation, recovery and recycling. Int J Hydrog Energy. 2019;44:3450–3455 https://doi.org/10.1016/j.ijhydene.2018.12.030. Cheng et al., 2009 Cheng S, Xing D, Call DF, et al. Direct biological conversion of electrical current into methane by electromethanogenesis. Environ Sci Technol. 2009;43:3953–3958 https://doi.org/10.1021/es803531g. Chen et al., 2014a Chen S, Rotaru AE, Liu F, et al. Carbon cloth stimulates direct interspecies electron transfer in syntrophic co-cultures. Bioresour Technol. 2014a;173:82–86 https://doi.org/10.1016/j.biortech.2014.09.009. Chen et al., 2014b Chen S, Rotaru AE, Shrestha PM, et al. Promoting interspecies electron transfer with biochar. Sci Rep. 2014b;4 https://doi.org/10.1038/srep05019. Cocco et al., 2014 Cocco R, Karri SBR, Knowlton T, eds. Introduction to fluidization. Chem Eng Prog. 2014;110:21–29.

Coutanceau et al., 2018 Coutanceau C, Baranton S, Audichon T. Chapter 3 – Hydrogen production from water electrolysis. In: Coutanceau C, Baranton S, Audichon TBT-HEP, eds. Hydrogen Energy and Fuel Cells Primers. Academic Press 2018;17–62. Available from: https://doi.org/10.1016/B978-0-12-8112502.00003-0. del Pilar Anzola Rojas et al., 2018 del Pilar Anzola Rojas M, Zaiat M, Gonzalez ER, De Wever H, Pant D. Effect of the electric supply interruption on a microbial electrosynthesis system converting inorganic carbon into acetate. Bioresour Technol. 2018;266:203–210 Available from: https://doi.org/10.1016/j.biortech.2018.06.074. Domigall, 2016 Domigall Y. Impact of power-to-gas storage on the integration of photovoltaic electricity into the grid: a simulation approach with distribution grid data of Switzerland. Comput Sci – Res Dev. 2016;31:9–15 https://doi.org/10.1007/s00450-014-0279-3. Eigenberger, 1992 Eigenberger G. Fixed bed reactors. Ullmann’s Encyclopedia of Industrial Chemistry. Vol. B4 Weinheirn: VCH VerJagsgesellschaft; 1992;199–238 . Eveloy, 2019 Eveloy V. Hybridization of solid oxide electrolysis-based powerto-methane with oxyfuel combustion and carbon dioxide utilization for energy storage. Renew Sustain Energy Rev. 2019;108:550–571 Available from: https://doi.org/10.1016/j.rser.2019.02.027. Eveloy and Gebreegziabher, 2018 Eveloy V, Gebreegziabher T. A review of projected power-to-gas deployment scenarios. Energies. 2018;11 https://doi.org/10.3390/en11071824. Fei et al., 2018 Fei F, Hui C, Lei Z, Liang J. Carbon deposition on nickel-based catalyst during bio-syngas methanation in a fluidized bed reactor. IOP Conf Ser Earth Environ Sci. 2018;199 https://doi.org/10.1088/1755-1315/199/3/032040. Feng et al., 2018 Feng Q, Song YC, Ahn Y. Electroactive microorganisms in bulk solution contribute significantly to methane production in bioelectrochemical anaerobic reactor. Bioresour Technol. 2018;259:119–127 https://doi.org/10.1016/j.biortech.2018.03.039. Feng et al., 2019 Feng F, Song G, Xiao J, Shen L, Pisupati SV. Carbon

deposition on Ni-based catalyst with TiO2 as additive during the syngas methanation process in a fluidized bed reactor. Fuel. 2019;235:85–91 https://doi.org/10.1016/j.fuel.2018.07.076. Fuentes et al., 2010 Fuentes RE, Farell J, Weidner JW. Multimetallic electrocatalysts of Pt, Ru, and Ir ed on anatase and rutile TiO2 for oxygen evolution in an acid environment. Electrochem Solid State Lett. 2010;14:E5. Furuya and Motoo, 1979 Furuya N, Motoo S. The electrochemical behavior of ad-atoms and their effect on hydrogen evolution Part IV Tin and lead ad-atoms on platinum. J Electroanal Chem Interf Electrochem. 1979;98:195–202 https://doi.org/10.1016/S0022-0728(79)80259-4. Ganigue et al., 2015 Ganigue R, Puig S, Batlle-Vilanova P, Balaguer MD, Colprim J. Microbial electrosynthesis of butyrate from carbon dioxide. Chem Commun. 2015;51:3235–3238 https://doi.org/10.1039/C4CC10121A. Geppert et al., 2019 Geppert F, Liu D, Weidner E, ter Heijne A. Redox-flow battery design for a methane-producing bioelectrochemical system. Int J Hydrog Energy. 2019;44:21464–21469 https://doi.org/10.1016/j.ijhydene.2019.06.189. Gorby et al., 2006 Gorby YA, Yanina S, McLean JS, et al. Electrically conductive bacterial nanowires produced by Shewanella oneidensis strain MR-1 and other microorganisms. Proc Natl Acad Sci. 2006;103:11358–11363 https://doi.org/10.1073/pnas.0905246106. Görke et al., 2005 Görke O, Pfeifer P, Schubert K. Highly selective methanation by the use of a microchannel reactor. Catal Today. 2005;110:132–139 https://doi.org/10.1016/j.cattod.2005.09.009. Götz et al., 2013 Götz M, Ortloff F, Reimert R, Basha O, Morsi BI, Kolb T. Evaluation of organic and ionic liquids for three-phase methanation and biogas purification processes. Energy Fuels. 2013;27:4705–4716 https://doi.org/10.1021/ef400334p. Götz et al., 2014 Götz, M., Koch, A.M.D., Graf, F., 2014. State of the art and perspectives of CO2 methanation process concepts for power-to-gas applications, In: Proceedings of the International Gas Union Research Conference. 1, pp. 314–327.

Götz et al., 2016 Götz M, Lefebvre J, Mörs F, et al. Renewable power-to-gas: a technological and economic review. Renew Energy. 2016;85:1371–1390 https://doi.org/10.1016/j.renene.2015.07.066. Grigoriev et al., 2006 Grigoriev SA, Porembsky VI, Fateev VN. Pure hydrogen production by PEM electrolysis for hydrogen energy. Int J Hydrog Energy. 2006;31:171–175 https://doi.org/10.1016/j.ijhydene.2005.04.038. Grubb, 1959a Grubb WT. Batteries with solid ion exchange electrolytes I Secondary cells employing metal electrodes. J Electrochem Soc. 1959a;106:275–278. Grubb, 1959b Grubb WT. Ionic migration in ion-exchange membranes. J Phys Chem. 1959b;63:55–58. Gruber et al., 2018 Gruber M, Wieland C, Habisreuther P, et al. Modeling and design of a catalytic wall reactor for the methanation of carbon dioxide. Chem Ing Tech. 2018;90:615–624. Guilera et al., 2020 Guilera J, Andreu T, Basset N, et al. Synthetic natural gas production from biogas in a waste water treatment plant. Renew Energy. 2020;146:1301–1308 https://doi.org/10.1016/j.renene.2019.07.044. Guo et al., 2013 Guo X, Liu J, Xiao B. Bioelectrochemical enhancement of hydrogen and methane production from the anaerobic digestion of sewage sludge in single-chamber membrane-free microbial electrolysis cells. Int J Hydrog Energy. 2013;38:1342–1347 https://doi.org/10.1016/j.ijhydene.2012.11.087. Guo et al., 2017 Guo Z, Thangavel S, Wang L, et al. Efficient methane production from beer wastewater in a membraneless microbial electrolysis cell with a stacked cathode: the effect of the cathode/anode ratio on bioenergy recovery. Energy Fuels. 2017;31:615–620 https://doi.org/10.1021/acs.energyfuels.6b02375. Hansen et al., 2016 Hansen, J.B., Holstebroe, M., Jensen, U.M.B., Rass-Hansen, J., Heiredal-Clausen, T., 2016. SOEC enabled biogas upgrading, In: Proceedings of the Twelfth European SOFC SOE Forum, pp. 1–10. Hashimoto et al., 2014 Hashimoto K, Kumagai N, Izumiya K, Takano H, Kato

Z. The production of renewable energy in the form of methane using electrolytic hydrogen generation. Energy Sustain Soc. 2014;4:17 https://doi.org/10.1186/s13705-014-0017-5. Hemmati et al., 2017 Hemmati R, Saboori H, Jirdehi MA. Stochastic planning and scheduling of energy storage systems for congestion management in electric power systems including renewable energy resources. Energy. 2017;133:380– 387 https://doi.org/10.1016/j.energy.2017.05.167. Januszewska et al., 2014 Januszewska A, Jurczakowski R, Kulesza PJ. CO2 electroreduction at bare and cu-decorated pd pseudomorphic layers: catalyst tuning by controlled and indirect ing onto Au(111). Langmuir. 2014;30:14314–14321 https://doi.org/10.1021/la5025247. Jiang et al., 2013 Jiang Y, Su M, Zhang Y, Zhan G, Tao Y, Li D. Bioelectrochemical systems for simultaneously production of methane and acetate from carbon dioxide at relatively high rate. Int J Hydrog Energy. 2013;38:3497–3502 https://doi.org/10.1016/j.ijhydene.2012.12.107. Jia et al., 2019 Jia C, Dai Y, Yang Y, Chew JW. Nickel–cobalt catalyst ed on TiO2-coated SiO2 spheres for CO2 methanation in a fluidized bed. Int J Hydrog Energy. 2019;44:13443–13455 https://doi.org/10.1016/j.ijhydene.2019.04.009. Jourdin et al., 2016 Jourdin L, Freguia S, Flexer V, Keller J. Bringing high-rate, CO2-based microbial electrosynthesis closer to practical implementation through improved electrode design and operating conditions. Environ Sci Technol. 2016;50:1982–1989 https://doi.org/10.1021/acs.est.5b04431. Jourdin et al., 2018 Jourdin L, Raes SMT, Buisman CJN, Strik DPBTB. Critical biofilm growth throughout unmodified carbon felts allows continuous bioelectrochemical chain elongation from CO2 up to caproate at high current density. Front Energy Res. 2018;6:1–15 https://doi.org/10.3389/fenrg.2018.00007. Khatib et al., 2019 Khatib FN, Wilberforce T, Ijaodola O, et al. Material degradation of components in polymer electrolyte membrane (PEM)electrolytic cell and mitigation mechanisms: a review. Renew Sustain Energy Rev. 2019;111:1–14 https://doi.org/10.1016/j.rser.2019.05.007.

Kopyscinski et al., 2010 Kopyscinski J, Schildhauer TJ, Biollaz SMA. Production of synthetic natural gas (SNG) from coal and dry biomass – a technology review from 1950 to 2009. Fuel. 2010;89:1763–1783 https://doi.org/10.1016/j.fuel.2010.01.027. Kreitz et al., 2019a Kreitz B, Friedland J, Güttel R, Wehinger GD, Turek T. Dynamic methanation of CO2–effect of concentration forcing. Chem Ingenieur Techn. 2019a;91:576–582 https://doi.org/10.1002/cite.201800191. Kreitz et al., 2019b Kreitz B, Wehinger GD, Turek T. Dynamic simulation of the CO2 methanation in a micro-structured fixed-bed reactor. Chem Eng Sci. 2019b;195:541–552 https://doi.org/10.1016/j.ces.2018.09.053. Lappas and Heracleous, 2011 Lappas A, Heracleous E. Biofuels from thermal and thermo-chemical conversion processes and technologies. In: Luque R, Campelo J, Clark JBT, eds. Handbook of Biofuels Production. Woodhead Publishing 2011;493–529. Available from: https://doi.org/10.1016/B978-184569-679-5.50026-7. Lefebvre, 2019 Lefebvre, J., 2019. Methanation Reaction Kinetics and Transient Behavior of a Slurry Bubble Column Reactor Three-Phase CO2 Methanation, p. 176. Lefebvre et al., 2018 Lefebvre J, Trudel N, Bajohr S, Kolb T. A study on threephase CO2 methanation reaction kinetics in a continuous stirred-tank slurry reactor. Fuel. 2018;217:151–159 https://doi.org/10.1016/j.fuel.2017.12.082. Liu et al., 2004 Liu H, Ramnarayanan R, Logan BE. Production of electricity during wastewater treatment using a single chamber microbial fuel cell. Environ Sci Technol. 2004;38:2281–2285 https://doi.org/10.1021/es034923g. Liu et al., 2005 Liu H, Grot S, Logan BE. Electrochemically assisted microbial production of hydrogen from acetate. Environ Sci Technol. 2005;39:4317–4320 https://doi.org/10.1021/es050244p. Liu et al., 2012 Liu Z, Chu B, Zhai X, Jin Y, Cheng Y. Total methanation of syngas to synthetic natural gas over Ni catalyst in a micro-channel reactor. Fuel. 2012;95:599–605 https://doi.org/10.1016/j.fuel.2011.12.045. Liu et al., 2015a Liu F, Rotaru AE, Shrestha PM, Malvankar NS, Nevin KP,

Lovley DR. Magnetite compensates for the lack of a pilin-associated c-type cytochrome in extracellular electron exchange. Environ Microbiol. 2015a;17:648–655 https://doi.org/10.1111/1462-2920.12485. Liu et al., 2015b Liu G, Xu J, Wang Y, Wang X. An oxygen evolution catalyst on an antimony doped tin oxide nanowire structured for proton exchange membrane liquid water electrolysis. J Mater Chem A. 2015b;3:20791–20800 https://doi.org/10.1039/C5TA02942B. Li et al., 2011 Li X, Walsh C, Pletcher D. Nickel based electrocatalysts for oxygen evolution in high current density, alkaline water electrolysers. Phys Chem Chem Phys. 2011;13:1162–1167 https://doi.org/10.1039/c000993h. Li et al., 2015 Li X, Abu-Reesh I, He Z. Development of bioelectrochemical systems to promote sustainable agriculture. Agriculture. 2015;5:367–388 https://doi.org/10.3390/agriculture5030367. Li et al., 2019 Li J, Agarwal RK, Zhou L, Yang B. Investigation of a bubbling fluidized bed methanation reactor by using CFD-DEM and approximate image processing method. Chem Eng Sci. 2019;207:1107–1120 Available from: https://doi.org/10.1016/j.ces.2019.07.016. Li et al., 2020 Li J, Agarwal RK, Yang B. Two-dimensional computational fluid dynamics simulation of heat removal in fluidized bed methanation reactors from coke oven gas using immersed horizontal tubes. Ind Eng Chem Res. 2020;59:981–991 https://doi.org/10.1021/acs.iecr.9b06557. Lovley et al., 2011 Lovley DR, Ueki T, Zhang T, et al. Geobacter: the microbe electric’s physiology, ecology, and practical applications. In: R.K.B.T.-A in Poole MP, ed. Advances in Microbial Physiology. Academic Press 2011;1–100. Available from: https://doi.org/10.1016/B978-0-12-387661-4.00004-5. Maksic et al., 2011 Maksic A, Brković S, Nikolic V, Perović I, Marceta Kaninski M. Energy consumption of the electrolytic hydrogen production using Ni–W based activators—part I. Appl Catal A Gen. 2011;405:25–28 https://doi.org/10.1016/j.apcata.2011.07.017. Malvankar et al., 2011 Malvankar NS, Vargas M, Nevin KP, et al. Tunable metallic-like conductivity in microbial nanowire networks. Nat Nanotechnol. 2011;6:573–579 https://doi.org/10.1038/nnano.2011.119.

Mamaca et al., 2012 Mamaca N, Mayousse E, Arrii-Clacens S, et al. Electrochemical activity of ruthenium and iridium based catalysts for oxygen evolution reaction. Appl Catal B Environ. 2012;111–112:376–380 Available from: https://doi.org/10.1016/j.apcatb.2011.10.020. Manabe et al., 2013 Manabe A, Kashiwase M, Hashimoto T, et al. Basic study of alkaline water electrolysis. Electrochim Acta. 2013;100:249–256 https://doi.org/10.1016/j.electacta.2012.12.105. Marceta Kaninski et al., 2011 Marceta Kaninski MP, Saponjic DP, Nikolic VM, Zugic DL, Tasic GS. Energy consumption and stability of the Ni–Mo electrodes for the alkaline hydrogen production at industrial conditions. Int J Hydrog Energy. 2011;36:8864–8868 Available from: https://doi.org/10.1016/j.ijhydene.2011.04.144. Marini et al., 2012 Marini S, Salvi P, Nelli P, et al. Advanced alkaline water electrolysis. Electrochim Acta. 2012;82:384–391 https://doi.org/10.1016/j.electacta.2012.05.011. Mathieu, 2010 Mathieu P. 9 - Oxyfuel combustion systems and technology for carbon dioxide (CO2) capture in power plants. In: M.M.B.T.-D. and I. in C.D. (CO2) C. and Maroto-Valer ST, ed. Developments and Innovation in Carbon Dioxide (CO2) Capture and Storage Technology, Woodhead Publishing Series in Energy. Woodhead Publishing 2010;283–319. Available from: https://doi.org/10.1533/9781845699574.3.283. Michishita et al., 2008 Michishita H, Matsumoto H, Ishihara T. Effects of pressure on the performance of water electrolysis of the cell using nafion membrane electrode. Electrochemistry. 2008;76:288–292 https://doi.org/10.5796/electrochemistry.76.288. Moreno et al., 2016 Moreno R, San-Martín MI, Escapa A, Morán A. Domestic wastewater treatment in parallel with methane production in a microbial electrolysis cell. Renew Energy. 2016;93:442–448 https://doi.org/10.1016/J.RENENE.2016.02.083. Morita et al., 2011 Morita M, Malvankar NS, Franks AE, et al. Potential for direct interspecies electron transfer in methanogenic wastewater digester aggregates. mBio. 2011;2:5–7 https://doi.org/10.1128/mBio.00159-11.

Mueller, 2012 Mueller, J.R., 2012. Microbial Catalysis of Methane From Carbon Dioxide, Ohio State University, pp. 1–107. Nam et al., 2020 Nam H, Kim JH, Kim H, et al. CO2 methanation in a benchscale bubbling fluidized bed reactor using Ni-based catalyst and its exothermic heat transfer analysis. Energy. 2020;214:118895 Available from: https://doi.org/10.1016/j.energy.2020.118895. Nelabhotla and Dinamarca, 2018 Nelabhotla ABT, Dinamarca C. Electrochemically mediated CO2 reduction for bio-methane production: a review. Rev Environ Sci Biotechnol. 2018;17:531–551 https://doi.org/10.1007/s11157-018-9470-5. Nelabhotla and Dinamarca, 2019 Nelabhotla ABT, Dinamarca C. Bioelectrochemical CO2 reduction to methane: MES integration in biogas production processes. Appl Sci. 2019;9:1056 https://doi.org/10.3390/app9061056. Nelabhotla et al., 2019 Nelabhotla ABT, Bakke R, Dinamarca C. Performance analysis of biocathode in bioelectrochemical CO2 reduction. Catalysts. 2019;9:683 https://doi.org/10.3390/catal9080683. Nevin et al., 2010 Nevin KP, Woodard TL, Franks AE. Microbial electrosynthesis: feeding microbial electrosynthesis: feeding microbes electricity to convert carbon dioxide and water to multicarbon extracellular organic. Am Soc Microbiol. 2010;1:1–4 https://doi.org/10.1128/mBio.00103-10.Editor. Nikolic et al., 2010 Nikolic VM, Tasic GS, Maksic AD, Saponjic DP, Miulovic SM, Marceta Kaninski MP. Raising efficiency of hydrogen generation from alkaline water electrolysis – energy saving. Int J Hydrog Energy. 2010;35:12369–12373 https://doi.org/10.1016/j.ijhydene.2010.08.069. Pestalozzi et al., 2019 Pestalozzi J, Bieling C, Scheer D, Kropp C. Integrating power-to-gas in the biogas value chain: analysis of stakeholder perception and risk governance requirements. Energy Sustain Soc. 2019;9 https://doi.org/10.1186/s13705-019-0220-5. Peters et al., 2019 Peters R, Baltruweit M, Grube T, Samsun RC, Stolten D. A techno economic analysis of the power to gas route. J CO2 Util. 2019;34:616– 634 https://doi.org/10.1016/j.jcou.2019.07.009.

Phillips and Dunnill, 2016 Phillips R, Dunnill CW. Zero gap alkaline electrolysis cell design for renewable energy storage as hydrogen gas. RSC Adv. 2016;6:100643–100651 https://doi.org/10.1039/C6RA22242K. Puthiyapura et al., 2013 Puthiyapura VK, Pasupathi S, Basu S, et al. RuxNb1− xO2 catalyst for the oxygen evolution reaction in proton exchange membrane water electrolysers. Int J Hydrog Energy. 2013;38:8605–8616. Rabaey and Rozendal, 2010 Rabaey K, Rozendal RA. Microbial electrosynthesis —revisiting the electrical route for microbial production. Nat Rev Microbiol. 2010;8:706–716 https://doi.org/10.1038/nrmicro2422. Raoof et al., 2010 Raoof JB, Ojani R, Esfeden SA, Nadimi SR. Fabrication of bimetallic Cu/Pt nanoparticles modified glassy carbon electrode and its catalytic activity toward hydrogen evolution reaction. Int J Hydrog Energy. 2010;35:3937–3944 Available from: https://doi.org/10.1016/j.ijhydene.2010.02.073. Rashid et al., 2015 Rashid MM, Al Mesfer MK, Naseem H, Danish M. Hydrogen production by water electrolysis: a review of alkaline water electrolysis, PEM water electrolysis and high temperature water electrolysis. Int J Eng Adv Technol. 2015;4:80–93. Reier et al., 2014 Reier T, Teschner D, Lunkenbein T, et al. Electrocatalytic oxygen evolution on iridium oxide: uncovering catalyst-substrate interactions and active iridium oxide species. J Electrochem Soc. 2014;161:F876–F882. Rönsch et al., 2017 Rönsch S, Ortwein A, Dietrich S. Start-and-stop operation of fixed-bed methanation reactors – results from modeling and simulation. Chem Eng Technol. 2017;40:2314–2321 https://doi.org/10.1002/ceat.201700229. Rosa et al., 1995 Rosa VM, Santos MBF, da Silva EP. New materials for water electrolysis diaphragms. Int J Hydrog Energy. 1995;20:697–700 Available from: https://doi.org/10.1016/0360-3199(94)00119-K. Rotaru et al., 2014 Rotaru A-E, Shrestha PM, Liu F, et al. A new model for electron flow during anaerobic digestion: direct interspecies electron transfer to Methanosaeta for the reduction of carbon dioxide to methane. Energy Environ Sci. 2014;7:408–415 https://doi.org/10.1039/C3EE42189A.

Rozendal et al., 2008 Rozendal RA, Jeremiasse AW, Hamelers HVM, Buisman CJNN. Hydrogen production with a microbial biocathode. Environ Sci Technol. 2008;42:629–634 https://doi.org/10.1021/Es071720+. Sarkar and Peter, 2018 Sarkar S, Peter SC. An overview on Pd-based electrocatalysts for the hydrogen evolution reaction. Inorg Chem Front. 2018;5:2060–2080 https://doi.org/10.1039/C8QI00042E. Sarno and Ponticorvo, 2019 Sarno M, Ponticorvo E. High hydrogen production rate on RuS2@ MoS2 hybrid nanocatalyst by PEM electrolysis. Int J Hydrog Energy. 2019;44:4398–4405. Schaaf et al., 2014 Schaaf T, Grünig J, Schuster MR, Rothenfluh T, Orth A. Methanation of CO2 - storage of renewable energy in a gas distribution system. Energy Sustain Soc. 2014;4:1–14 https://doi.org/10.1186/s13705-014-0029-1. Schiller et al., 2019 Schiller G, Lang M, Szabo P, et al. Solar heat integrated solid oxide steam electrolysis for highly efficient hydrogen production. J Power Sources. 2019;416:72–78 https://doi.org/10.1016/j.jpowsour.2019.01.059. Seemann and Thunman, 2019 Seemann M, Thunman H. 9-Methane synthesis. In: Materazzi M, ed. Substitute Natural Gas From Waste. Academic Press 2019;221–243. Available from: https://doi.org/10.1016/B978-0-12-8155547.00009-X. Shiva Kumar and Himabindu, 2019 Shiva Kumar S, Himabindu V. Hydrogen production by PEM water electrolysis – a review. Mater Sci Energy Technol. 2019;2:442–454 https://doi.org/10.1016/j.mset.2019.03.002. Shrestha et al., 2014 Shrestha PM, Malvankar NS, Werner JJ, et al. Correlation between microbial community and granule conductivity in anaerobic bioreactors for brewery wastewater treatment. Bioresour Technol. 2014;174:306–310 https://doi.org/10.1016/j.biortech.2014.10.004. Siegert et al., 2014 Siegert M, Yates MD, Call DF, Zhu X, Spormann A, Logan BE. Comparison of nonprecious metal cathode materials for methane production by electromethanogenesis. ACS Sustain Chem Eng. 2014;2:910–917 https://doi.org/10.1021/sc400520x. Specht et al., 2010 Specht M, Brellochs J, Frick V, et al. Speicherung von

Bioenergie und erneuerbarem Strom im Erdgasnetz, Erdöl, Erdgas, Kohle. 2010;126:342. Stojić et al., 2003 Stojić D, Marceta Kaninski M, Sovilj S, Miljanić Š. Hydrogen generation from water electrolysis – possibilities of energy saving. J Power Sources. 2003;118:315–319 https://doi.org/10.1016/S0378-7753(03)00077-6. Sun et al., 2018 Sun L, Luo K, Fan J. Production of synthetic natural gas by CO methanation over Ni/Al2O3 catalyst in fluidized bed reactor. Catal Commun. 2018;105:37–42 https://doi.org/10.1016/j.catcom.2017.11.003. Takenaka, 1991 Takenaka H. Development trends of hydrogen production technology by water electrolysis. J Fuel Soc Jpn. 1991;70:487–496 https://doi.org/10.3775/jie.70.487. Tasic et al., 2011 Tasic GS, Maslovara SP, Zugic DL, Maksic AD, Marceta Kaninski MP. Characterization of the Ni–Mo catalyst formed in situ during hydrogen generation from alkaline water electrolysis. Int J Hydrog Energy. 2011;36:11588–11595 Available from: https://doi.org/10.1016/j.ijhydene.2011.06.081. Thema et al., 2019 Thema M, Bauer F, Sterner M. Power-to-gas: electrolysis and methanation status review. Renew Sustain Energy Rev. 2019;112:775–787 https://doi.org/10.1016/j.rser.2019.06.030. Thiele et al., 1988 Thiele JH, Chartrain M, Zeikus JG. Control of interspecies electron flow during anaerobic digestion: role of floc formation in syntrophic methanogenesis. Appl Environ Microbiol. 1988;54:10–19 https://doi.org/10.1128/aem.54.1.10-19.1988. Uhm and Kim, 2014 Uhm S, Kim YD. Electrochemical conversion of carbon dioxide in a solid oxide electrolysis cell. Curr Appl Phys. 2014;14:672–679 https://doi.org/10.1016/j.cap.2014.02.013. Ursúa et al., 2012 Ursúa A, Gandía LM, Sanchis P. Hydrogen production from water electrolysis: current status and future trends. Proc IEEE. 2012;100:410– 426 https://doi.org/10.1109/JPROC.2011.2156750. van Eerten-Jansen et al., 2015 van Eerten-Jansen MCAA, Jansen NC, Plugge CM, de Wilde V, Buisman CJN, ter Heijne A. Analysis of the mechanisms of

bioelectrochemical methane production by mixed cultures. J Chem Technol Biotechnol. 2015;90:963–970 https://doi.org/10.1002/jctb.4413. Wang et al., 2019a Wang L, Chen M, Küngas R, et al. Power-to-fuels via solidoxide electrolyzer: operating window and techno-economics. Renew Sustain Energy Rev. 2019a;110:174–187 Available from: https://doi.org/10.1016/j.rser.2019.04.071. Wang et al., 2019b Wang L, Düll J, Maréchal F, Van herle J. Trade-off designs and comparative exergy evaluation of solid-oxide electrolyzer based power-tomethane plants. Int J Hydrog Energy. 2019b;44:9529–9543 Available from: https://doi.org/10.1016/j.ijhydene.2018.11.151. Witte et al., 2019 Witte J, Calbry-Muzyka A, Wieseler T, Hottinger P, Biollaz SMA, Schildhauer TJ. Demonstrating direct methanation of real biogas in a fluidised bed reactor. Appl Energy. 2019;240:359–371 https://doi.org/10.1016/j.apenergy.2019.01.230. Xu et al., 2006 Xu G, Chen X, Zhang Z-G. Temperature-staged methanation: an alternative method to purify hydrogen-rich fuel gas for PEFC. Chem Eng J. 2006;121:97–107 https://doi.org/10.1016/J.CEJ.2006.05.010. Xu et al., 2012 Xu J, Liu G, Li J, Wang X. The electrocatalytic properties of an IrO2/SnO2 catalyst using SnO2 as a and an assisting reagent for the oxygen evolution reaction. Electrochim Acta. 2012;59:105–112 Available from: https://doi.org/10.1016/j.electacta.2011.10.044. Xu et al., 2014 Xu H, Wang K, Holmes DE. Bioelectrochemical removal of carbon dioxide (CO2): an innovative method for biogas upgrading. Bioresour Technol. 2014;173:392–398 https://doi.org/10.1016/j.biortech.2014.09.127. Yan and Hino, 2016 Yan XL, Hino R. Nuclear Hydrogen Production Handbook CRC Press 2016. Yeh et al., 2018 Yeh M-H, Leu Y-A, Chiang W-H, et al. Boron-doped carbon nanotubes as metal-free electrocatalyst for dye-sensitized solar cells: Heteroatom doping level effect on tri-iodide reduction reaction. J Power Sources. 2018;375:29–36 Available from: https://doi.org/10.1016/j.jpowsour.2017.11.041. Yekini Suberu et al., 2014 Yekini Suberu M, Wazir Mustafa M, Bashir N. Energy

storage systems for renewable energy power sector integration and mitigation of intermittency. Renew Sustain Energy Rev. 2014;35:499–514 https://doi.org/10.1016/j.rser.2014.04.009. Yin et al., 2016 Yin Q, Zhu X, Zhan G, et al. Enhanced methane production in an anaerobic digestion and microbial electrolysis cell coupled system with cocultivation of Geobacter and Methanosarcina. J Environ Sci (Chin.). 2016;42:210–214 https://doi.org/10.1016/j.jes.2015.07.006. Zahid et al., 2010 Zahid M, Schefold J, Brisse A. High-temperature water electrolysis using planar solid oxide fuel cell technology: a review. Hydrogen and Fuel Cells: Fundamentals, Technologies and Applications Weinheim: WileyVCH; 2010;227–242. Zhang et al., 2020 Zhang, H., Wang, L., Van herle, J., Maréchal, F., Desideri, U., 2020. Techno-economic evaluation of biomass-to-fuels with solid-oxide electrolyzer, Appl. Energy. 270, 115113. Available from: https://doi.org/10.1016/j.apenergy.2020.115113. Zhao et al., 2014 Zhao Z, Zhang Y, Chen S, Quan X, Yu Q. Bioelectrochemical enhancement of anaerobic methanogenesis for high organic load rate wastewater treatment in a up-flow anaerobic sludge blanket (UASB) reactor. Sci Rep. 2014;4:6658–6665 https://doi.org/10.1038/srep06658. Zhao et al., 2015 Zhao Z, Zhang Y, Wang L, Quan X. Potential for direct interspecies electron transfer in an electric-anaerobic system to increase methane production from sludge digestion. Sci Rep. 2015;5:11094 https://doi.org/10.1038/srep11094. Zhao et al., 2016 Zhao Z, zhang Y, Li Y, Zhao H, Quan X. Electrochemical reduction of carbon dioxide to formate with Fe-C electrodes in anaerobic sludge digestion process. Water Res. 2016;106:339–343 https://doi.org/10.1016/j.watres.2016.10.018. Zhao et al., 2017 Zhao Z, Zhang Y, Li Y, Dang Y, Zhu T, Quan X. Potentially shifting from interspecies hydrogen transfer to direct interspecies electron transfer for syntrophic metabolism to resist acidic impact with conductive carbon cloth. Chem Eng J. 2017;313:10–18 https://doi.org/10.1016/j.cej.2016.11.149.

Chapter 9

Electrochemical approach for biogas upgrading

Grzegorz Pasternak, Laboratory of Microbial Electrochemical Systems, Department of Process Engineering and Technology of Polymer and Carbon Materials, Wroclaw University of Science and Technology, Wrocław, Poland

Abstract

An increasing fossil fuel combustion raise the environmental concerns in of environmental pollution and climate change. Biogas is a byproduct of microbial metabolism. It is a sustainable fuel made of organic waste such as biomass and may reduce the global demand in fossil fuels. Its methane (CH4) content usually reaches 50%–70%, while its second most abundant component is carbon dioxide (CO2). In order to increase the caloric value of biogas, the methane content needs to be increased through an upgrading process. In this chapter, a possibility of using electrochemical methods to upgrade biogas by removal of its two major problematic compounds will be discussed. This chapter will focus on electroreduction of carbon dioxide (CO2), which is the second major compound in biogas, as well as electrochemical oxidation of hydrogen sulfide (H2S), which is its most corrosive component. Basic fundamental aspects, electrocatalysts, electrolyzer, and fuel cell designs, as well as challenges for further development and industrial applications toward electrochemical biogas upgrading will be discussed.

Keywords

CO2; H2S; reduction; oxidation; electrolysis; biogas; upgrading; fuel cell

Chapter outline

Outline


9.2 Faradaic and energy efficiency 226

9.3 Electroreduction of CO2226

9.3.1 Basic considerations 227 9.3.2 Reactor and process design 230

9.4 Electrochemical oxidation of H2S 236

9.4.1 Basic considerations 237 9.4.2 Reactor and process design 237

9.5 Biogas upgrading approach and its challenges 241

9.5.1 CO2 electroreduction 241 9.5.2 H2S oxidation 243 9.5.3 Biogas and scale-up approaches 244

9.6 Concluding remarks and perspectives 246

Acknowledgments 247

References 247

9.1 Introduction

In recent decades, climate change concerns have been constantly growing due to increasing average temperatures recorded around the world. One of the major sources of atmospheric pollution with greenhouse gases is the combustion of fossil fuels (Ritchie and Roser, 2020; Pasternak et al., 2011). Meantime, massive reserves of carbon are produced and stored in the form of solid organic waste, resulting from anthropogenic activity. These wastes can be converted into biogas, which leads to a decrease in the demand for natural gas extraction. Biogas is a valuable byproduct of anaerobic microbial metabolism. Microbial production of biogas is based on waste resources such as wastewater or municipal waste. Such an environment is reach in substrates and reveals high microbial diversity. The microorganisms are supplied with a variety of metabolic pathways and the resulting biogas is a mixture of compounds, in which the methane (CH4) may reach a concentration of up to 80%. Carbon dioxide (CO2) remains as the second major component of biogas, on average reaching 30%– 50% of its total volume, while the other components are nitrogen (N2), water (H2O), hydrogen sulfide (H2S), ammonia (NH3), chlorides (Cl−), hydrocarbons, and in some cases, also siloxanes (Angelidaki et al., 2018; Awe et al., 2017; Chen et al., 2015). All of these compounds decrease the caloric value of the biogas. While the main compound responsible for decreased caloric value of biogas is CO2, some of the compounds may cause serious risks to the operation of the technological process such as toxicity and corrosion caused by H2S and NH3, or accumulation of silicone oxide deposits. Therefore, their presence may cause damage in the engines and other mechanical parts during combustion. Following the above, biogas requires upgrading—a process in which the impurities are removed and the methane concentration is increased. The conversion of waste and wastewater into valuable products meets the goals of a circular economy and has attracted the attention of scientists and industry worldwide. Biogas production has been recognized as a potential, significant source of energy by multiple countries, of which , Sweden, and the United Kingdom have developed nearly 70% of worldwide biogas plants (Angelidaki et al., 2018). Several methods have been applied to upgrade biogas,

with water scrubbers the most commonly used so far. In water scrubbers, CO2 is absorbed by water in which it has a higher absorption rate than methane. Although simple, the technology requires significant amounts of water reaching up to 200 m³/h (Ryckebosch et al., 2011). The process is often accompanied by the Claus process, in which the H2S is removed through thermal and subsequent catalytic conversion. The absorption rate of CO2 in chemical solvents is higher than water. Nevertheless, chemical solvents such as commercial organic solvents or amine solutions are difficult to regenerate and toxic for the environment, which along with their costs are their main drawbacks (Persson, 2003; Adnan et al., 2019). Lower consumption of chemicals is achieved when sodium hydroxide is used as the absorbent (Angelidaki et al., 2018; Yoo et al., 2013). Another technique relies on physical separation of biogas contaminants using membrane processes, where CH4 permeation is much slower than N2, H2S, CO2, and water, in that order. The process requires pretreatment steps: H2S removal to avoid corrosion and H2O removal to prevent the loss of efficiency followed by the compression to 5–20 bars (Bauer et al., 2020), and is considered as costly when compared to chemical and physical scrubbers (Adnan et al., 2019). Substantial attention in recent years has been given to a separate group of techniques, which rely on biological processes for biogas upgrading. Biological reactors offer the unique opportunity of combining multiple reactions into one platform process. Embedding microorganisms in such a synthetic environment results in extending their outstanding capabilities. Two principal methods in of metabolic pathways are used: chemoautotrophy and photoautotrophy. Chemoautotrophic methods rely on hydrogenotrophic methanogens which produce CH4 from CO2 and require H2 to perform the process (Maegaard et al., 2019). The H2 can be delivered from renewable energy sources through the water-splitting electrochemical techniques (Dau et al., 2010) or through bioelectrochemical techniques. Bioelectrochemical systems are of particular interest due to their sustainability and circular economy potential (Pasternak et al., 2020; Pasternak et al., 2019). These techniques rely on a variant of microbial fuel cells, namely microbial electrolysis cells (MEC), where bacteria produce electrons and protons from chemical compounds at the anode which are further electrochemically combined at the cathode to produce hydrogen (Logan et al., 2008; Rousseau et al., 2020). When CO2 reduction takes place in such a system, the MEC can produce methane at the cathode (Nelabhotla et al., 2019). Another sustainable way of consuming CO2 from biogas is the use of microalgae species in a photoautotrophic process (del R Rodero et al., 2020). The microalgal growth requires either a natural or artificial light source. The biological processes are

carried out in mild operational conditions, however their use in biogas upgrading is relatively new when compared to other methods and requires further work toward industrialization. All of the aforementioned methods have their strengths and weaknesses, which are discussed in detail within other chapters of this book. Herein, another interesting—yet still under development—group of techniques will be discussed, which comprise of electrochemical means of reducing the amount of undesirable compounds in biogas. Electrochemical reduction and oxidation of compounds has gained significant attention in recent years as a promising method to convert renewable resources such as CO2 or H2S into value-added products and to recover these resources from waste streams. Such an approach may be further developed toward biogas upgrading, where the concentration of biogas impurities may be reduced, thus increasing its caloric value. In this chapter, this growing field and a review of the fundamental and technological aspects will be discussed, as well as the most recent progress of using electrochemical methods toward biogas upgrading. The electrochemical conversion methods are growing rapidly and a variety of simple and complex organic and inorganic compounds have been investigated so far. This chapter, however, will be focused solely on the two most important components of biogas: CO2, which is the second major compound in biogas after CH4, and H2S, which is the major corrosive and inhibiting compound for industrial storage, transportation, and processing infrastructure. In the case of the conversion of other biogas components, in particular when considering nitrogen oxidation and nitrogen reduction reaction, the reader is advised to refer to some of the most recent, and comprehensive, reviews available in the literature (Dai et al., 2020; Suryanto et al., 2019; Xue et al., 2019; Guo et al., 2019).

9.2 Faradaic and energy efficiency

The electrochemical techniques have two major and comparative parameters that describe their efficiency, which are Faradaic (Eq. 9.1) and energy (Eq. 9.2) efficiencies. The Faradaic efficiency is a measure of efficiency in which the electrons are contributing to a desired electrochemical reaction, while energy efficiency determines the overall energy consumption of the process. Using and reporting these parameters allows comparing different approaches and technologies as well as estimating their cost efficiency.

(9.1)

(9.2)

where z is the number of electrons exchanged for the product, n is the number of moles of the product, F is the Faradaic constant (F=96.485 C/mol), Q is the total charge applied (C), Et is the theoretical potential for the formation of the product (V), and E is the actual applied potential (V).

9.3 Electroreduction of CO2

The electrochemical reduction, also referred to as electroreduction of CO2, is labeled as CO2R. The process may take place either in low- or high-temperature conditions, and combinations of systems in gaseous, aqueous, and nonaqueous conditions (Bevilacqua et al., 2015). Low-temperature processes (0°C–100°C) mainly rely on the use of transition metal electrodes as catalysts in CO2/H2Oliquid electrolyte systems. High-temperature processes (500°C–1000°C) are mainly based on solid-oxide fuel cell (SOFCs)—derivatives called solid-oxide electrolysis cells (SOECs) and are operated in CO2 or CO2/H2O steam conditions (Spinner et al., 2012). The high-temperature CO2R leads to the formation of H2, CO, and CH4, while low-temperature CO2R additionally produces organic C1 and C2 compounds (Fig. 9.1). In principle, such systems may be also used to generate hydrogen and to use it for the subsequent methanation of CO2. The major advantage of SOECs and SOFCs over liquid electrolyte devices is their higher selectivity toward the end products, while the latter offers the advantage of running the process in mild temperature conditions which makes them more feasible for distributed energy systems. Numerous excellent reviews on CO2R are available in the literature, covering broad aspects of the process from the fundamentals to applications (Adnan et al., 2019; Bevilacqua et al., 2015; Hori, 2008; Jhong et al., 2013; Nitopi et al., 2019; Kortlever et al., 2015b; Sánchez et al., 2019; Liang et al., 2020; Whang et al., 2019; Wang et al., 2017; Chen et al., 2018). Thus, in the subsequent sections, only the major considerations of the most exploited aspects of the CO2R process and its focus on biogas upgrading will be presented.

Figure 9.1 Commonly reported pathways for two major processes of electrochemical conversion of CO2. The equilibrium potentials are given for each product versus reversible hydrogen electrode (RHE) calculated in standard thermodynamic temperature (25°C). Each final CO2R product requires protons derived from a water-splitting process (oxygen evolution reaction, OER), thus 1.23 V adds up to the overall equation, while the cell voltage sums anodic, cathodic potential and the overpotential (difference between theoretical and actual potential of the electrode caused by various potential losses). The number of electrons for low- and high-temperature pathways for the most common products (H2, CH4, and CO) are the same and thus are represented only once for clarity. Based on data from Spinner N.S., Vega J.A., Mustain W.E., 2012. Recent progress in the electrochemical conversion and utilization of CO2. Catal. Sci. Technol. 2, 19–28. https://doi.org/10.1039/c1cy00314c and Nitopi, S., Bertheussen, E., Scott, S.B., Liu, X., Engstfeld, A.K., Horch, S., et al., 2019. Progress and perspectives of electrochemical CO2 reduction on copper in aqueous electrolyte. Chem. Rev. 119, 7610–7672. https://doi.org/10.1021/acs.chemrev.8b00705.

9.3.1 Basic considerations

In the CO2R process, the CO2 is converted to a variety of products according to Eq. (9.3), while CO may be further reduced (Eq. 9.4) and such a process is designated as COR. Both reactions are cathodic:

(9.3)

(9.4)

Following Eqs. (9.1) and (9.2), the CO2R process is considered as a multiplestep reaction, where protons and electrons are used to generate various products, along with water. The equilibrium potentials for various reactions have been given in Fig. 9.1. In order for the process to be sustainable, the CO2R reaction (Eq. 9.3) needs to be balanced by the oxygen evolution reaction (OER) resulting from electrochemical water splitting (Dau et al., 2010), where the hydrogen evolution reaction (HER) also occurs according to the half-reactions:

(9.5)

(9.6)

The occurrence of HER in the CO2R process is undesirable since HER competes with the electrochemical reduction of CO2 and decreases the Faradaic and energy efficiency of the process. Similarly, it has been reported that the presence of oxygen in the cathodic CO2R process may compete with it and significantly reduce its Faradaic efficiency (Xu et al., 2020).

9.3.1.1 Solid-oxide devices

The high-temperature conversion of CO2 is most commonly carried out in SOECs and leads to the production of syngas (CO/H2). The obtained syngas

may be further used in the production of various synthetic fuels and products (Wilhelm et al., 2001). Hydrogen as the final product may be also used for subsequent hydrogenation of CO2 to methane and thus to further upgrade biogas. In contrast to SOFCs, where the energy is being produced, SOECs require energy input (Fig. 9.2), which should be delivered from renewable sources to meet the profitability and sustainability objectives. Both types of devices have been reported to electrochemically convert CO2 either to valuable products (SOECs) or to electricity (SOFCs) (Ebbesen and Mogensen, 2009; Huang and Chou, 2009), although in the majority of conventional SOFCs, the CO2 is an end product of fuel oxidation.

Figure 9.2 The difference in operating principles of a solid-oxide electrolysis cell and a solid-oxide fuel cell. Reprinted from Ebbesen, S.D., Mogensen, M., 2009. Electrolysis of carbon dioxide in solid oxide electrolysis cells. J. Power Sources 193, 349–358. https://doi.org/10.1016/j.jpowsour.2009.02.093, with permission from Elsevier.

The SOECs operate at between 300°C and 1500°C (Chen et al., 2018), which significantly reduces their electrical energy demand, as well as the charge transfer and transport overpotentials for the cathodic CO2R reaction (Eq. 9.7), where co-electrolysis with water (Eq. 9.8) is also possible (Bevilacqua et al., 2015). The solid-oxide electrolyzers require catalysts, for which the nickel-based materials are the most commonly used (Whang et al., 2019). CO2 electrolysis:

(9.7)

Steam electrolysis:

(9.8)

Reverse water-gas shift reaction (RWGS):

(9.9)

Oxygen formation:

(9.10)

Another possible pathway for CO2R has been reported for proton-conducting SOECs, where protons diffuse through a conducting electrolyte to the cathode, where the methane can be formed (Eq. 9.11), while competing with RWGS (Eq. 9.9) and hydrogen generation (Eq. 9.6) (Xie et al., 2011):

(9.11)

The drawback of solid-oxide devices in of biogas upgrading comes from the fact that CH4 may be also converted simultaneously to produce hydrogen in the electroreduction process. Furthermore, the presence of CH4 may also lead to electrode deactivation through its decomposition and deposition of coke (Kim et al., 2002).

9.3.1.2 Liquid electrolyte devices

Liquid-phase electroreduction takes place in mild temperature (usually below 60°C) and pressure conditions. Similarly as for solid-oxide electrolyzers, CO2 is

reduced at the cathode, where the reaction competes with hydrogen evolution. The most conventional experimental setup for studying the CO2R in electrolyte systems is an H-type cell, presented in Fig. 9.3. In order for the reduction to take place at low temperatures, the presence of a catalyst is required. The catalysts are transition metals and metal oxides. A variety of metals have been reported (Kortlever et al., 2015b; Whang et al., 2019), among which the most promising have been recognized as the copper-based electrodes, for which a very comprehensive review has been recently published (Nitopi et al., 2019). A great variety of products have been also reported and their simplified pathways are illustrated in Fig. 9.1. The actual pathways to produce a desired product are, in fact, very complex and dependent on the type of electrocatalyst and therefore will not be discussed herein. Instead, the reader is referred to other published works where this topic has been elaborated on (Nitopi et al., 2019; Kortlever et al., 2015b; Handoko et al., 2018). The variety of possible electrochemical pathways and their overlapping potentials leads to the poor selectivity of the process. It is difficult to obtain a desired, specific product of choice, as well as to selectively convert the CO2 from the mixture of gases. This poor selectivity is mainly caused by the reactivity of the CH4 and the overlapping potential ranges of CO2 with other biogas components.

Figure 9.3 Simplified representation of the most common designs for studying the electroreduction of CO2 in (A) solid-oxide electrolysis cells and (B) liquid electrolyte systems, as well as continuous-flow electrochemical reactors: (C) a liquid electrolyzer where CO2 is dissolved in electrolyte, (D) CO2 is delivered in the gaseous phase, with the anode submerged in water, and (E) a solid-oxide electrolysis flow cell. Reprinted from (A) Wang, Y., Liu, T., Lei, L., Chen, F., 2017. High temperature solid oxide H2O/CO2 coelectrolysis for syngas production. Fuel Process. Technol. https://doi.org/10.1016/j.fuproc.2016.08.009, (B) Zhao, C., Wang, J., 2016. Electrochemical reduction of CO2 to formate in aqueous solution using electro-deposited Sn catalysts. Chem. Eng. J. 293, 161–170. https://doi.org/10.1016/j.cej.2016.02.084, (C)–(E) Chen, C., Kotyk, J.F.K., Sheehan, S.W., 2018. Progress toward commercial application of electrochemical carbon dioxide reduction. Chem 4, 2571–2586. https://doi.org/10.1016/j.chempr.2018.08.019, with permission from Elsevier.

9.3.2 Reactor and process design

The electrochemical reduction of CO2 is performed in a variety of designs. Nevertheless, the two most commonly used designs have been established for electrolysis (Fig. 9.3A and B). The SOEC in principle consists of two electrodes, separated by the solid ion-conducting electrolyte (Fig. 9.3A). The electrodes are porous, most commonly ceramic composite materials made of nickel and yttriastabilized zirconia (YSZ). The solid electrolyte can be either oxygen ionconducting or proton-conducting. In the first case, water and CO2 are reduced in the cathode and oxygen ions to the anode where they are oxidized to O2. In the latter one, CO2 is reduced at the cathode, while water is oxidized to O2 and protons travel to the cathode to react with CO2 and form products. Since the heat reduces external resistance, high-temperature SOECs are of particular interest, where due to increased temperature, precious metal catalysts may be avoided and generated heat can be re-used. The H-type cell, used for electrolyte-based reactions (Fig. 9.3B) consists of the

cathodic and anodic chambers, separated with the semi-permeable membrane. The anodic chamber, where the water splitting takes place, is usually equipped with the Pt electrode, while the cathodic chamber is supplied with the electrocatalyst-electrode for the CO2 reduction. The CO2 is pumped to the cathode chamber and the outlet is directed to gas analysis, while electrolyte from both chambers is analyzed for the presence of products. The system is equipped with the reference electrode and potentiostat to control its electrochemical regime. So far this system has been recognized as a commercially available labscale setup (Liang et al., 2020). The most commonly used membranes so far are cation (CEM) and anion (AEM) exchange membranes. The vast majority of reports focus on carrying out the reduction process in alkaline electrolyte, among which potassium hydrogen carbonate, potassium, and sodium hydroxides are the most commonly used. An increased pH leads to a shift in the CO2 equilibrium toward CO3²−, which reduces the availability of the dissolved CO2 in the electrolyte. Nevertheless, alkaline conditions favor CO2R over competing HER which occurs at low pH (König et al., 2019). Depending on the design, choice of electrolyte, and experimental conditions, different types of products and Faradaic efficiencies can be observed. Fundamental works on design have revealed that reducing the spacing between the electrodes resulted in lowering the ohmic resistance of the cell and improving its performance (Liang et al., 2020; Kuhl et al., 2012). Eventually this concept evolved as a modification of the H-cell design, where the planar electrodes and membrane are sandwiched, and is thus referred to as a sandwichtype. Carrying out the process in batch conditions, as provided in the H-type cells, has some critical drawbacks like a pH imbalance resulting from CO2 depletion (thus affecting the carbonate equilibrium), as well as the mass-transfer issue and product accumulation, which in turn affect the resistance of the system. To overcome these issues, continuous-flow cells are often investigated (Fig. 9.3C–F), either in their sandwich-type or microfluidic variants. Such an approach also facilitates product collection and allows to establish a stable and reliable electrosynthesis. Over recent decades, multiple architectures have been proposed for the continuous CO2R, while the majority of them can be classified according to their feeding conditions as presented in Fig. 9.3. The vast majority of these works are currently focused toward investigations of various electrocatalysts, the substrate for their deposition, type, and composition of the electrolyte, temperature, pressure, and the membrane properties such as permselectivity. When deg the CO2R process toward biogas upgrading,

the operating pressure seems to play a crucial role as affects the activity and selectivity of the electrocatalyst. According to Bevilacqua et al. (2015), the metal catalysts can be grouped into four groups, depending on the way the pressure affects their activity:

1. Low activity in atmospheric and high pressures (Ti, Nb, Ta, Mo, Mn, AI); 2. Low activity at atmospheric pressure but considerable activity at higher pressures (Zr, Cr, W, Fe, Co, Rh, Ir, Ni, Pd, Pt, C, Si); 3. Elements able to convert CO2 into formate at low and high pressures (Ag, Au, Zn, In, Sn, Pb Bi); 4. Copper (Cu), which is able to convert CO2 to various products, depending on the applied pressure.

For the SOECs, the most established, robust anodic material with high stability and catalytic activity is Ni/YSZ composite cermet (Shri Prakash et al., 2014). Similarly, Ni-based cermet composites are used as cathode catalysts. The H2S present in biogas may significantly reduce the catalytic activity of such materials (Hauch et al., 2014). Thus, particular focus is also given to reducing such risk. As an example, in the work of Park et al. (2019), the authors developed a material based on CoNi alloy nanoparticles anchored on a Ruddlesden–Popper (RP) of La1.2Sr0.8Co0.4Mn0.6O4. Such a catalyst revealed stable performance without a decrease or deterioration in the material after 90 h of operation under H2S presence and exposure with up to 100 ppm H2S/N2, while also reaching a maximum FE of 97.8% and current density of 703 mA cm². Another important aspect of the electrode design comprises of its mass-transfer efficiency of substrates and products to and from the electrode’s surface. Therefore several research groups have been depositing metals on the diffusion layer (GDL, also GDE—gas diffusion electrode), which facilitates mass transport within a multiphase environment (Sánchez et al., 2019; Bajracharya et al., 2016). Depending on the approach for catalyst synthesis, electrode structure, and design architecture, a variety of efficiencies have been reported. In the most recent report, Liang et al. (2020) gathered the Faradaic efficiencies of the CO2R

process recorded by several research groups for various experimental designs (Table 9.1).

Table 9.1

Reactor type H-type cell

Electrocatalyst Conventional H-type cell

Substrate A-Ni-NG (Yang et al., 2018) NS-C (Pan et al., 2019) Np-Ag (Lu et al., 2014) OD-Cu (Kas et al., 2014) Cu nanocube (Jiang et al., 2018) PdPt/C (Kortlever et al., 2015a)

Sandwich-type cell

Cu foil (Kuhl et al., 2012) Cu NPs (Manthiram et al., 2014) AN-Cu (Lee et al., 2018) Ag foil (Hatsukade et al., 2014)

Pressurized reactor Flow cell

Cu NPs (Kas et al., 2015) PEM

In/Pb (Narayanan et al., 2011)

Pt/C, Pd/C, Cu/CNTs (Wang et al., 2018)

Ag (Salvatore et al., 2018)

CoPc (Ren et al., 2019)

Pd, Ag, and Zn (Lee et al., 2019)

Cu oxides/ZnO (Merino-Garcia et al., 2019)

Cu2O/ZnO (Albo et al., 2017)

Microfluidic

RuPd/Sn (Whipple et al., 2010) Ag (Rosen et al., 2011) CuNP (Ma et al., 2016) Graphite/carbon NPs/Cu/PTFE (Dinh et al., 2018)

SOEC

La0.80Sr0.20Sc0.05Mn0.95O3-d - Gd0.20Ce0.80O1.95 (Yu Ni/YSZ (Ebbesen et al., 2012)

DEMS cell

Cu sheet (Clark et al., 2015) Cu/Ag (Clark and Bell, 2018)

Source: Reprinted from Liang, S., Altaf, N., Huang, L., Gao, Y., Wang, Q., 2020. Electrolytic cell design for electrochemical CO2 reduction, J. CO2 Util. 35, 90– 105. https://doi.org/10.1016/j.jcou.2019.09.007 with permission from Elsevier.

According to this summary, the highest Faradaic efficiencies can be observed when CO is the target product of the CO2R process. An efficiency as high as 97%, along with a low overpotential of −0.61 V versus RHE, was obtained by Yang et al. (2018) using the atomic dispersion technique, where nickel was dispersed on nitrogenated graphene. Promising results for conversion of CO2 to CH4 were also obtained by Manthiram et al. (2014), who supplied a sandwichtype cell (a more compact variant of the H-cell) with a nanoparticle copper catalyst dispersed on the surface of glassy carbon. The authors claimed to achieve up to four times higher methanation current densities as compared to high-purity copper foil electrodes, and very high values of FE reaching average values of 80%. Interesting results were also reported by Kas et al. (2015), who altered the pressure conditions and electrolyte concentration which resulted in changing the pH near the electrode, surface. In such an environment, various ratios of methane and ethylene were observed, where ethylene formation was predominantly formed at a high local pH. Although in theory this could be a good approach for increasing the CH4 amount in the biogas upgrading process, the authors also recorded a significant deterioration in further product formation when CH4 was produced, while the formation of ethylene maintained the electrocatalytic activity of the electrode for an extended time. Ethylene was also synthesized from CO2 by Merino-Garcia et al., who developed Cu oxides/ZnObased electrodes, which revealed excellent FE equal to 91.1% and a production rate of 487.9±37 µmol/m²/s. The authors claimed that using a synergistic effect of Cu and Zn favors the selectivity of the process toward production of ethylene, although this was only true when the Cu nanoparticles as a catalyst were not included in the comparison. For such a material the authors recorded a higher production rate of 1148±136 µmol/m²/s and the selectivity toward C2H4/H2 was 50% higher when compared to Cu oxides/ZnO (Merino-Garcia et al., 2018).

9.4 Electrochemical oxidation of H2S

Hydrogen sulfide is known to cause corrosion of steel and costly economical losses within the oil and gas industries (Wiener et al., 2006). Moreover, as already mentioned it may severely affect the CO2R process. Its removal is therefore necessary in of biogas upgrading. The electrochemical means of hydrogen sulfide conversion have been an object of interest since early 20th century, when Fetzer showed a possibility of electrochemical oxidation of sodium sulfide (Fetzer, 1928). The subsequent studies showed that it is possible to oxidize H2S and recover two valuable products through H2S electrolysis: elemental sulfur and hydrogen (Mao, 1991; Zaman and Chakma, 1995). Nevertheless, the problem associated with elemental sulfur deposition on the electrode has not yet been successfully solved. Such a deposition leads to termination of the whole process.

9.4.1 Basic considerations

Hydrogen sulfide is a weak acid and in aqueous solutions dissociates to H+ + HS − or 2H+ + S²−. It is also susceptible to oxidation, which leads to the formation of sulfates and elemental sulfur (Xu et al., 2016). At the electrode’s surface, direct oxidation of H2S gas is accompanied by a release of two electrons, according to Eq. (9.12).

(9.12)

In general, H2S oxidation occurs in two types of electrochemical reactions: direct, when the compound is susceptible to electrochemical reactions, and indirect with the involvement of mediators or oxidants, which takes place in bulk electrolyte (Pikaar et al., 2015). A direct sulfide oxidation may lead to the formation of other sulfur ions such as thiosulfate (S2O3²−), sulfite (SO3²−), and sulfate (SO4²−). From an economical point of view, the preferred pathway is the conversion to elemental sulfur (2e−) in contrast to sulfate (8e−). A low potential of direct oxidation of H2S to sulfur is the main advantage of the process, which favors its decomposition when compared to water. H2S oxidation is accompanied by HER occurring at the cathode (Eq. 9.6). The electrochemical pathways involved in indirect oxidations are dependent on the method. Different methods for indirect oxidation may involve Ag²+, Ce⁴+, Fe²+, and (VO2)+ ions (Pikaar et al., 2015; Narendranath et al., 2019; Mejia Likosova et al., 2014). Indirect oxidation is usually a non-selective process, since the reactive species diffuse to the electrolyte and may oxidize a wide range of other compounds. In most studies, the electrolysis process is carried out in alkaline conditions, where in the preliminary step, H2S is dissolved in an alkaline solution, leading to the formation of sulfide ion (Eq. 9.13). Anodic oxidation (Eq. 9.14) leads to the formation of sulfur, which may dissolve in alkaline media to form polysulfides, while the water reduction takes place at the cathode (Eq. 9.15) to form hydrogen gas and hydroxide ions. In such a system, the gas containing H2S needs to be sparged through the alkaline media.

(9.13)

(9.14)

(9.15)

9.4.2 Reactor and process design

In principle, the electrolytic device design is similar to those which are used for liquid electrolysis in CO2R (Fig. 9.3B), which involves a membrane between the electrode compartments, with the difference that the reaction of interest takes place at the anode and requires energy input. An example of such a design specifically for H2S splitting is shown in Fig. 9.4A. Such a system has been used in most studies from the past decades both for direct and indirect oxidation (Mao, 1991; Narendranath et al., 2019; Selvaraj et al., 2016; Guo et al., 2015; Alexander and Winnick, 1994). A similar design, though as a fuel cell concept, has been also proposed (Fig. 9.3B), where the H2S was used as a source of electrons at the anode, and potassium ferricyanide was used as an electron acceptor at the cathode, thus leading to the production of electricity (Dutta et al., 2008). Furthermore, membraneless systems have been investigated for H2S electrooxidation including real waste (Fig. 9.3C) (Wang et al., 2019). The family of solid-oxide devices, which could be also a potential technology for H2S oxidation (Fig. 9.4D), is severely affected by the presence of H2S in the system, which is corrosive for their conventional anodic materials (Choi et al., 2019; Sun et al., 2018). Therefore, recent research has rather focused toward sulfur-tolerant materials than mechanisms and applications of H2S oxidation (Aguilar et al., 2004). Another family of hybrid technologies such as photoelectrochemical (Qiao et al., 2018) methods will not be discussed herein. Several up-to-date reviews have been recently published in this area (Pawar et al., 2019; Xu and Carter, 2019; Castro et al., 2018).

Figure 9.4 Electrochemical splitting of the H2S process: (A) direct alkaline electrolysis as proposed by Ntagia et al. (2020) and (B) electrochemical oxidation in a fuel cell with the aid of ferricyanide reduction by Dutta et al. (2008). (C) A single-chamber reactor for H2S oxidation in swine manure by Wang et al. (2019). (D) SOFC operating on H2S by Liu et al. (2001). (A) Adapted and (B)–(D) reprinted with permission from Elsevier.

The experiments on H2S removal have been carried out both in gaseous phase and electrolyte, including aqueous and nonaqueous systems. In gaseous phase systems the gas containing H2S is usually ed through the scrubber, which absorbs H2S in liquid and further reactions of such electrolyte take place in an electrochemical reactor. Selvaraj et al. (2016) investigated the recovery of elemental sulfur from gas ed through an NaOH scrubber, using a reactor assembly with titanium substrate insoluble anode meshes and the influence of their geometry on process efficiency at 20 mA/cm. The initial current was, however, reduced by more than 50% for all of the geometries after approximately 4 h from the start of the experiments due to sulfur deposition. In another study, Narendranath et al. (2019) used two electrode architectures based on RuO2-TiO2. Similarly to the previous studies (Mao, 1991), the process was found to be optimal at 80°C and high alkaline conditions of pH 13. The authors observed 91% sulfur recovery. Although the experiments were carried out in continuous mode and electrodes of perforated geometry were displayed as free of sulfur deposits, the long-term performance of the electrodes was not demonstrated. Similarly to the CO2R process, the electrode is a critical component of the electrochemical system. Apart from its geometry, its functionality is also developed toward new types of materials and surface modifications, while in the past the most common catalyst was Pt (Szynkarczuk et al., 1994). Mbah et al. (2010) used a nanostructured composite catalyst built from RuO2-CoS2. The reactor was running at 135 kPa and 150°C and treating pure H2S gas. Stable operation was demonstrated for a period of 24 h, reaching a current density of 19 mA/cm² and 69% conversion efficiency, these results were correlated with an active surface area against the other tested electrode materials. Interestingly, in one of the other studies, where a larger reactor was built for the process (Gerçel et al., 2008), the Pt-based mesh cathodes were determined to be the main catalyst

responsible for the H2S conversion to polysulfides, while at the same time, the authors sparged H2S through an alkaline electrolyte and the electrochemical mechanism of possible cathodic conversion was not explained. An alternative approach for indirect H2S oxidation was proposed by Guo et al. (2015) as a non aqueous system, where imidazolium chloride ionic liquid (Fe(III)-IL) was used as a desulfurizer and combined with Fe(II) and N,N-dimethylformamide (DMF). A resulting non aqueous system [Fe(III/II)-IL/DMF] was regenerated through the application of potential at its optimal value of 0.800 V (vs Ag/AgCl). Following up to six 5-h cycles, the system remained stable in of removal efficiency, which reached 98%. In contrast to the electrolysis reactors, Dutta et al. (2008) showed that electrochemical oxidation of H2S may take place in a spontaneous manner, with concomitant generation of electricity, and thus using the H-type design as a fuel cell and turning H2S as a source of electricity (Fig. 9.4B). The current was generated with the aid of ferricyanide, which was reduced at the cathode, while the H2S was oxidized at the cathode. Interestingly, the system was able to operate for 2 months in continuous-flow conditions at an average rate of H2S reduction of 0.62 kg S/m³/d and the average power density generated was 12±2 W/m³. A significant decay of performance caused by the sulfur deposition was observed after 3 months. When compared to the other studies, where the H2S oxidation was usually stable for a period of hours (Selvaraj et al., 2016; Mbah et al., 2010), this may be considered as a good result. Nevertheless, the rate of reaction was relatively low and using ferricyanide is considered as problematic due to its possible HCN release in acidic conditions. In the case of SOFC devices, some recent works have focused on developing anodic materials capable of tolerating H2S, which even in trace amounts may entirely deactivate conventional materials based on Ni, Pt, and Ag (Aguilar et al., 2004), and cause either reversible or irreversible deterioration of the anodes (Hauch et al., 2014). The mechanism leading to performance deterioration is based on physical blocking of sulfur deposits of the electrode active sites or chemical reaction with nickel sulfide at the anode (Xu et al., 2010). Such a deterioration, for example, was observed by Liu et al. (2001) who developed the system based on Pt/(ZrO2)0.92(Y2O3)0.08/Pt electrode/solid electrolyte. At 5% H2S concentration, atmospheric pressure and temperature of 800°C, the SOFCs reached 15.4 mW/cm² performance. The formation of PtS however, resulted in detachment of the Pt anode from the YSZ membrane and corresponding deterioration of performance. Several approaches have been undertaken to

address this issue and new, sulfur-tolerant materials have been developed. Aguilar et al. (2004) have proposed LaxSr1−xVO3−δ (LSV) anodes for H2Scontaining fuels. The LSV materials appeared to be more selective toward H2S oxidation in comparison to H2 in a mixture of gases—a feature which could be also useful in a biogas upgrading process. An entire setup in LSV/YSZ/LSMYSZ configuration displayed the power density of 90 mW/cm² at 220 mA/cm² at 5% H2S in N2 and 135 mW/cm² at 280 mA/cm² at 95% H2S–5% H2. The stable performance was demonstrated for 48 h of operation while a long-term scenario for H2S poisoning in more recent studies was investigated for 600 h (Hauch et al., 2014). The durability of the electrodes must certainly be extended in order for H2S electro-oxidation to be applied at an industrial scale.

9.5 Biogas upgrading approach and its challenges

9.5.1 CO2 electroreduction

The economic feasibility of various CO2R products needs to be taken into when considering the scaling-up process. Such an analysis, performed by Chen et al. (2018), showed that, in 2017, only the electricity costs calculated for CO2 conversion to CH4 would exceed its current market value, although this analysis did not include the indirect value of turning the rest of the CH4 already present in biogas into useful product through upgrading. Among the CO2R products, formic acid was revealed to be the most promising one, reaching the highest value of revenue per mole of electrons consumed, equal to 0.015 USD and followed by CO with approximately 0.009 USD/mole e−. Several other techno economic studies have been carried out, among which the most recent work of Na et al. (2019) indicated that a further decrease of the operating costs of CO2R may be achieved through the oxidation of organic chemicals instead of water. The authors simulated 295 electrochemical coproduction processes within over 130,000 simulations and included biomass intermediates in the analysis. The results revealed that 2-furoic acid and FDCA obtained in the anodic organic oxidation reaction coupled with the CO2R process can add up to the profitability of cathodic synthesis of various chemicals. Such results indicate that an electrochemical coproduction, where the residual biomass from the biogas production is being oxidized in the anodic reaction, can be envisioned and truly approach circular economy assumptions. Although tremendous amount of works on CO2R have been reported in recent decades, the commercialization of the technology is still not yet implemented and therefore biogas upgrading through electrochemical means remains only within the laboratory scale. There are several technological obstacles in order to reach that goal. In particular, CO2 electrolysis is highly susceptible to the impurities which are the components of the real feedstocks, such as biogas, as well as industrial or atmospheric gas. Therefore the stability of the process is

limited, as is the electrode and membrane lifetime. Most studies have focused on the hours of lifetime of the catalysts. Each of the electrolyzer designs which have been listed in the previous section have their significant limitations which need to be addressed prior to scaling-up the process. A comparison of these features, as well as scale-up approaches, has been recently revised by Sánchez et al. (2019). As shown in Table 9.2 the CO2 electrolyzers are still in the development phase, where the functional elements like electrodes and membranes are not fully understood and require more studies. This statement in particular applies to long-term stability and durability of the materials in the presence of impurities, as well as their contribution to the design aspects which affect the overall resistance of the system and its efficiency.

Table 9.2

Reactor configuration Batch reactors

Continuous-flow reactors

Advantages

Drawbacks

One-compartment cells

• Simple design • Convenient for small-scale ex

Two-compartment cells (H-cell)

• Convenient to study half-cell reactions • Conv

Membrane reactor

• Relatively easy to upscale • Different configur

Microfluidic reactor

• No membrane included • Fast screening of cat

Solid-oxide electrolyzer

• Decrease in overpotentials • Increase in charge

Source: Reprinted after Sánchez, O.G., Birdja, Y.Y., Bulut, M., Vaes, J., Breugelmans, T., Pant, D., 2019. Recent advances in industrial CO2 electroreduction. Curr. Opin. Green Sustain. Chem. 16, 47–56. https://doi.org/10.1016/j.cogsc.2019.01.005, with permission from Elsevier.

Another important aspect is understanding the selectivity of electrocatalysts and their efficiency in the presence of impurities. An important example is the influence of oxygen on biogas, which results in an oxygen reduction reaction (ORR) as a side reaction which may completely inhibit the CO2 reduction, depending on the molar ratio of O2/CO2 gas. This issue has been recently addressed by Xu et al. (2020), who developed the oxygen-tolerant hydrated ionomer catalyst coating over the electrode. The electrode was thus made of hydrophilic nanopores bound with TiO2 nanoparticles at the top of the Cu catalyst. The ionomer was able to reduce the O2 mass transport rate, and the FE of CO2R to C2 compounds reached 68%. Another approach to overcome the effects of biogas components on CO2R stability, selectivity, and performance would require additional preliminary steps, such as separating CO2 from biogas or removing H2S from the system. The latter, however, could be also performed with the use of electrochemical methods. These are the main factors for which the industrial-scale applications of CO2R have not yet been reached (Sánchez et al., 2019; Chen et al., 2018).

9.5.2 H2S oxidation

Similarly to CO2R, only a few approaches have been undertaken to directly study H2S behavior in electrochemical devices fueled with biogas. Such an approach was studied by Xu et al. (2010), who operated SOFCs with synthetic biogas made from various ratios of pure gases. A 20 ppm H2S addition to the

mixture into Ni-YSZ-based SOFC resulted in immediate (2 h) reaction inhibition and both electrochemical and mechanical failures were observed. Modifying anodes with CeO2 resulted in only a slight improvement in durability. A realbiogas-operated SOFC in another study revealed that a concentration range from 1 to 2 ppm H2S had already inhibiting effects on electrochemical oxidation of biogas during 20–50 h trials (Shiratori et al., 2008; Papurello and Lanzini, 2018). It is worth noting that biogas may contain up to 10,000 ppm (Angelidaki et al., 2018) of H2S, which may clearly inhibit the electrocatalytic activity of the electrodes and even exceeds the requirements of H2S removal through adsorption processes. These examples show the principal struggles which need to be combated prior to implementation of the electrochemical methods toward H2S removal from biogas. These aspects so far have not been fully addressed either in SOFCs or low-temperature, liquid electrolyzers. Furthermore, an electrochemical process of H2S removal already possesses strong and wellestablished competitors such as wet absorption followed by the Claus process, where part of the H2S is burned to SO2 which subsequently reacts with residual H2S to form elemental sulfur (Pikaar et al., 2015).

9.5.3 Biogas and scale-up approaches

Regardless of the obstacles mentioned above, some actual biogas approaches, scaled-up systems, as well as systems combined with other technologies, have already been tested toward biogas treatment. Many of these studies focused rather on using biogas as a fuel and source of electricity rather than upgrading techniques to remove the biogas impurities. A combined approach, where the biogas is firstly pretreated and then treated in SOFCs was investigated by Chiodo et al. (2015). An outcome of the initial reforming process was syngas which was subsequently used as a fuel for SOFCs. Among steam reforming (biogas reforming with the use of water steam), autothermal reforming (where part of the biogas is combusted along with steam and oxygen addition), and partial oxidation (where part of the biogas is oxidized) as the preliminary step, steam reforming appeared to be the most efficient (DC electrical efficiency of 60%). A H2S removal step was added prior to SOFCs in pilot-scale, long-term studies, where biogas from a methane fermentation tank (Tosu Kankyo Kaihatsu

Ltd., Tosu, Japan) was used as a fuel, while H2O was removed by cold-trapping at 0°C. A semistable (due to biogas composition) voltage was observed during 1 month of operation, although severe coking was also noticed (Shiratori et al., 2010). An efficient desulfurization step toward <0.1 ppm was found to be critical to avoid coking, as well as mixing biogas with oxygen, which resulted in predominant H2 and CO oxidation. In contrast to using biogas as a fuel, an interesting example of the use of electrochemical reduction/oxidation processes specifically for biogas upgrading was recently proposed by Verbeeck et al. (2019). The lab-scale electrochemical reactor based on alkaline catholyte was proposed as a part of the upgrading system where the second step of (further) upgrading consisted of a biological reactor (Fig. 9.5). A synthetic biogas was fed to the cathodic chamber of an electrolyzer, where water splitting resulted in increasing the pH and absorbing H2S (HS−) and CO2 (HCO3−). The ions traveled to the anodic chamber through the AEM where the sulfide oxidation occurred and the CO2 was removed by stripping. The system required relatively high potentials of 2.51 to 6.47 V which corresponded to 3.73 to 17.38 kWh/Nm³ of CO2 removed. The observed H2S removal efficiencies ranged from 98%±1% to 100%±1% leaving 160 to 180 ppmv of H2S at the outlet, correspondingly. An economic feasibility of the process coupled with biological reactor was estimated at 0.72 €/Nm³ of raw biogas. In this proof-of-concept study, the trials comprised of a 3-h experiment, but the long-term performance was not investigated.

Figure 9.5 An electrochemical pretreatment approach for biogas upgrading. Hydrogen sulfide is removed through AEM and CO2 is treated in subsequent biomethanation reactor. MFC in this figure stands for mass flow controller, while PS for power supply. Reprinted from Verbeeck, K., De Vrieze, J., Biesemans, M., Rabaey, K., 2019. Membrane electrolysis-assisted CO2 and H2S extraction as innovative pretreatment method for biological biogas upgrading. Chem. Eng. J. 361, 1479–1486. https://doi.org/10.1016/j.cej.2018.09.120, with permission from Elsevier.

Apart from the actual experimental studies, biogas upgrading was also an objective of theoretical and modeling studies to display its feasibility (Gattrell et al., 2007). Several technological pipeline configurations were investigated through simulations of SOEC by Jeanmonod et al. (2019). Multiple concepts were presented and the results indicated that direct electrolysis of biogas in SOEC was the least preferred method of biogas upgrading, while the coelectrolysis of CO2 and steam accompanied by pretreatment steps was the most preferred method in of conversion efficiencies. Nevertheless, the economic feasibility of the proposed variants was not investigated. Such an analysis was carried out recently (Lorenzi et al., 2017) and the lowest cost of synthetic natural gas production from raw biogas was estimated to be 4.87 c€/kWhexergy, which positioned the investigated variants of electrolysis process as noneconomical. So far, the demonstration of direct biogas feeding into an electrolyzer at pilot-scale was only demonstrated in the largest existing power-to-gas (PtG) plant—Audi egas (). The plant consists of three, 2 MW electrolyzers although a smaller, alpha plant of 25 kW (ETOGAS) was used to demonstrate that in situ methanation is possible in such a system (Bailera et al., 2017). Another indirect approach for electrochemical biogas upgrading, which was not discussed in this chapter, comprises of hydrogen production in SOEC and subsequent methanation of CO2, which was demonstrated by Haldor Topsøe company (Denmark) at 50 kW SOEC scale which resulted in upgrading biogas from 56% to 97.69% (Hansen et al., 2016).

9.6 Concluding remarks and perspectives

This chapter reviews recent developments of electrochemical methods toward the reduction and oxidation of two important components of biogas which are critical for its upgrading: CO2 as a major component of biogas after CH4 and H2S as the major corrosive biogas component. The latter, along with siloxanes, chlorides, and ammonia strongly inhibits the electrocatalytic and mechanical durability of electrolyzers and SOFC devices. The SOFCs may be particularly sensitive to siloxanes such as octamethylcyclotetrasiloxane (D4) in trace amounts at the ppbv level (Papurello and Lanzini, 2018), while electrolysis cells carrying out CO2 reduction and H2S electrooxidation will suffer from electrode deactivation. The electrochemical oxidation of H2S occurs at very low potential, which would make this technique very attractive for H2S removal. Unfortunately, the major challenge, which is electrode deactivation due to sulfur deposition and chemical reaction with electrocatalysts, has not yet been solved. The CO2R process offers great value in a sustainable approach to obtain valueadded products and recirculate anthropogenically released carbon. Combining it with biogas upgrading would increase the environmental benefit. Nevertheless, the process struggles with the formation of coke at the surface of the electrodes in solid-oxide electrolyzers, and this aspect remains one of its major technological challenges (Bevilacqua et al., 2015). Considering low-temperature electrolyzers, an important consideration which is currently preventing successful scale-up of the technology is comprised of poor selectivity of the process and the electrochemical competition with H2 evolution which takes place at the same potential range. Furthermore, the O2 presence in biogas competes at the cathode with the CO2 reduction—instead, ORR occurs (Xu et al., 2020). These factors are accompanied by high overpotentials which further affect the profitability of the CO2R (Kortlever et al., 2015b). A scarcity of industrial or even pilot-scale studies for the direct electrochemical conversion methods of CO2 and H2S (Sánchez et al., 2019) demonstrates the scale of the challenge. Although a lot of fundamental aspects of these processes have been investigated and a number of new improved materials are being

described each year, the principal challenges still require further work. Future directions in industrialization of the electrochemical processes should be now focused toward long-term durability studies, accompanied with technoeconomical analyses. In parallel, materials which are tolerant of such biogas impurities as sulfur, siloxanes, and oxygen should be also developed. Since development of novel materials will require further fundamental studies, particular attention should be given to indirect methods where electrochemistry still plays the main role but the entire pipeline is built from several existing methods. In such hybrid technologies, combining both advanced separation materials and well-established separation processes as preliminary steps may be a technologically feasible way toward biogas electro-upgrading. An example of such an approach was demonstrated in ETOGAS—the Audi e-gas plant where the CO2 for the CO2R process was derived from a chemical scrubber. Finally, the closest path to implementing electrochemical methods for biogas upgrading may lead through integrating more conventional, stand-alone water electrolysis process with subsequent methanation of the CO2.

Acknowledgments

This work was ed by the Polish National Agency for Academic Exchange – Polish Returns grant (PPN/PPO/2018/1/00038); National Science Centre (Poland) OPUS grant (2019/33/B/NZ9/02774) and a subsidy from the Polish Ministry of Science and Higher Education for the Faculty of Chemistry of Wrocław University of Science and Technology. The author would also like to acknowledge the reviewers for their valuable input.

References

Adnan et al., 2019 Adnan AI, Ong MY, Nomanbhay S, Chew KW, Show PL. Technologies for biogas upgrading to biomethane: a review. Bioengineering. 2019;6:1–23 https://doi.org/10.3390/bioengineering6040092. Aguilar et al., 2004 Aguilar L, Zha S, Cheng Z, Winnick J, Liu M. A solid oxide fuel cell operating on hydrogen sulfide (H2S) and sulfur-containing fuels. J Power Sources 2004; https://doi.org/10.1016/j.jpowsour.2004.03.061. Albo et al., 2017 Albo J, Beobide G, Castaño P, Irabien A. Methanol electrosynthesis from CO2 at Cu2O/ZnO prompted by pyridine-based aqueous solutions. J CO2 Util. 2017;18:164–172 https://doi.org/10.1016/J.JCOU.2017.02.003. Alexander and Winnick, 1994 Alexander SR, Winnick J. Removal of hydrogen sulfide from natural gas through an electrochemical membrane separator. AIChE J. 1994;40:613–620 https://doi.org/10.1002/aic.690400406. Angelidaki et al., 2018 Angelidaki I, Treu L, Tsapekos P, et al. Biogas upgrading and utilization: current status and perspectives. Biotechnol Adv. 2018;36:452– 466 https://doi.org/10.1016/j.biotechadv.2018.01.011. Awe et al., 2017 Awe OW, Zhao Y, Nzihou A, et al. A review of biogas utilisation, purification and upgrading technologies. Waste Biomass Valor. 2017;8:267–283. Bailera et al., 2017 Bailera M, Lisbona P, Romeo LM, Espatolero S. Power to Gas projects review: lab, pilot and demo plants for storing renewable energy and CO2. Renew Sustain Energy Rev. 2017;69:292–312 https://doi.org/10.1016/j.rser.2016.11.130. Bajracharya et al., 2016 Bajracharya S, Vanbroekhoven K, Buisman CJN, Pant D, Strik DPBTB. Application of gas diffusion biocathode in microbial electrosynthesis from carbon dioxide. Environ Sci Pollut Res 2016;

https://doi.org/10.1007/s11356-016-7196-x. Bauer, et al., 2013 Bauer, F., Hulteberg, C., Persson, T., Tamm D., 2013. Biogas upgrading – review of commercial technologies. www.sgc.se. (accessed 13.07.20). Bevilacqua et al., 2015 Bevilacqua M, Filippi J, Miller HA, Vizza F. Recent technological progress in CO2 electroreduction to fuels and energy carriers in aqueous environments. Energy Technol. 2015;3:197–210 https://doi.org/10.1002/ente.201402166. Castro et al., 2018 Castro S, Albo J, Irabien A. Photoelectrochemical reactors for CO2 Utilization. ACS Sustain Chem Eng 2018; https://doi.org/10.1021/acssuschemeng.8b03706. Chen et al., 2015 Chen XY, Vinh-Thang H, Ramirez AA, Rodrigue D, Kaliaguine S. Membrane gas separation technologies for biogas upgrading. RSC Adv 2015; https://doi.org/10.1039/c5ra00666j. Chen et al., 2018 Chen C, Kotyk JFK, Sheehan SW. Progress toward commercial application of electrochemical carbon dioxide reduction. Chem. 2018;4:2571– 2586 https://doi.org/10.1016/j.chempr.2018.08.019. Chiodo et al., 2015 Chiodo V, Galvagno A, Lanzini A, et al. Biogas reforming process investigation for SOFC application. Energy Convers Manage 2015; https://doi.org/10.1016/j.enconman.2015.03.113. Choi et al., 2019 Choi DW, Ohashi M, Lozano CA, Vanzee JW, Aungkavattana P, Shimpalee S. Sulfur diffusion of hydrogen sulfide contaminants to cathode in a micro-tubular solid oxide fuel cell. Electrochim Acta 2019; https://doi.org/10.1016/j.electacta.2019.134713. Clark and Bell, 2018 Clark EL, Bell AT. Direct observation of the local reaction environment during the electrochemical reduction of CO2. J Am Chem Soc 2018; https://doi.org/10.1021/jacs.8b04058. Clark et al., 2015 Clark EL, Singh MR, Kwon Y, Bell AT. Differential electrochemical mass spectrometer cell design for online quantification of products produced during electrochemical reduction of CO2. Anal Chem 2015; https://doi.org/10.1021/acs.analchem.5b02080.

Dai et al., 2020 Dai C, Sun Y, Chen G, Fisher AC, Xu ZJ. Electrochemical oxidation of nitrogen towards direct nitrate production on spinel oxides. Angew Chem Int Ed 2020; https://doi.org/10.1002/anie.202002923. Dau et al., 2010 Dau H, Limberg C, Reier T, Risch M, Roggan S, Strasser P. The mechanism of water oxidation: from electrolysis via homogeneous to biological catalysis. Chem Cat Chem. 2010;2:724–761 https://doi.org/10.1002/cctc.201000126. del R Rodero et al., 2020 del R Rodero M, Carvajal A, Arbib Z, et al. Performance evaluation of a control strategy for photosynthetic biogas upgrading in a semi-industrial scale photobioreactor. Bioresour Technol 2020; https://doi.org/10.1016/j.biortech.2020.123207. Dinh et al., 2018 Dinh CT, Burdyny T, Kibria G, et al. CO2 electroreduction to ethylene via hydroxide-mediated copper catalysis at an abrupt interface. Science 2018; https://doi.org/10.1126/science.aas9100. Dutta et al., 2008 Dutta PK, Rabaey K, Yuan Z, Keller J. Spontaneous electrochemical removal of aqueous sulfide. Water Res. 2008;42:4965–4975 https://doi.org/10.1016/j.watres.2008.09.007. Ebbesen and Mogensen, 2009 Ebbesen SD, Mogensen M. Electrolysis of carbon dioxide in solid oxide electrolysis cells. J Power Sources. 2009;193:349–358 https://doi.org/10.1016/j.jpowsour.2009.02.093. Ebbesen et al., 2012 Ebbesen SD, Knibbe R, Mogensen M. Co-electrolysis of steam and carbon dioxide in solid oxide cells. J Electrochem Soc 2012; https://doi.org/10.1149/2.076208jes. Fetzer, 1928 Fetzer WR. The electrolysis of sodium sulphide solutions. J Phys Chem. 1928;32:1787–1807 https://doi.org/10.1021/j150294a002. Gattrell et al., 2007 Gattrell M, Gupta N, Co A. Electrochemical reduction of CO2 to hydrocarbons to store renewable electrical energy and upgrade biogas. Energy Convers Manage. 2007;48:1255–1265 https://doi.org/10.1016/j.enconman.2006.09.019. Gerçel et al., 2008 Gerçel Ö, Koparal AS, Ögˇütveren ÜB. Removal of hydrogen sulfide by electrochemical method with a batchwise operation. Sep Purif

Technol. 2008;62:654–658 https://doi.org/10.1016/j.seppur.2008.03.019. Guo et al., 2015 Guo Z, Zhang T, Liu T, et al. Nonaqueous system of iron-based ionic liquid and DMF for the oxidation of hydrogen sulfide and regeneration by electrolysis. Environ Sci Technol. 2015;49:5697–5703 https://doi.org/10.1021/es505728f. Guo et al., 2019 Guo W, Zhang K, Liang Z, Zou R, Xu Q. Electrochemical nitrogen fixation and utilization: theories, advanced catalyst materials and system design. Chem Soc Rev 2019; https://doi.org/10.1039/c9cs00159j. Handoko et al., 2018 Handoko AD, Wei F, Jenndy, Yeo BS, Seh ZW. Understanding heterogeneous electrocatalytic carbon dioxide reduction through operando techniques. Nat Catal. 2018;1:922–934 https://doi.org/10.1038/s41929018-0182-6. Hansen, et al., 2016 Hansen, J.B., Holstebroe, M., Jensen, U.M.B., Rass-Hansen, J., Heiredal-Clausen, T., 2016. SOEC enabled biogas upgrading. 12th Eur. SOFC SOE Forum 2016. 1–10. Hatsukade et al., 2014 Hatsukade T, Kuhl KP, Cave ER, Abram DN, Jaramillo TF. Insights into the electrocatalytic reduction of CO2 on metallic silver surfaces. Phys Chem Chem Phys 2014; https://doi.org/10.1039/c400692e. Hauch et al., 2014 Hauch A, Hagen A, Hjelm J, Ramos T. Sulfur poisoning of SOFC anodes: effect of overpotential on long-term degradation. J Electrochem Soc 2014; https://doi.org/10.1149/2.080406jes. Hori, 2008 Hori Y. Electrochemical CO2 reduction on metal electrodes. Mod Asp Electrochem 2008; https://doi.org/10.1007/978-0-387-49489-0_3. Huang and Chou, 2009 Huang TJ, Chou CL. Electrochemical CO2 reduction with power generation in SOFCs with Cu-added LSCF-GDC cathode. Electrochem Commun 2009; https://doi.org/10.1016/j.elecom.2009.05.031. Jeanmonod et al., 2019 Jeanmonod G, Wang L, Diethelm S, Maréchal F, Van herle J. Trade-off designs of power-to-methane systems via solid-oxide electrolyzer and the application to biogas upgrading. Appl Energy 2019; https://doi.org/10.1016/j.apenergy.2019.04.055.

Jhong et al., 2013 Jhong HRM, Ma S, Kenis PJ. Electrochemical conversion of CO2 to useful chemicals: current status, remaining challenges, and future opportunities. Curr Opin Chem Eng. 2013;2:191–199 https://doi.org/10.1016/j.coche.2013.03.005. Jiang et al., 2018 Jiang K, Sandberg RB, Akey AJ, et al. Metal ion cycling of Cu foil for selective C-C coupling in electrochemical CO2 reduction. Nat Catal 2018; https://doi.org/10.1038/s41929-017-0009-x. Kas et al., 2014 Kas R, Kortlever R, Milbrat A, Koper MTM, Mul G, Baltrusaitis J. Electrochemical CO2 reduction on Cu2O-derived copper nanoparticles: controlling the catalytic selectivity of hydrocarbons. Phys Chem Chem Phys 2014; https://doi.org/10.1039/c401520g. Kas et al., 2015 Kas R, Kortlever R, Yilmaz H, Koper MTM, Mul G. Manipulating the hydrocarbon selectivity of copper nanoparticles in CO2 electroreduction by process conditions. ChemElectroChem 2015; https://doi.org/10.1002/celc.201402373. Kim et al., 2002 Kim T, Moon S, Hong SI. Internal carbon dioxide reforming by methane over Ni-YSZ-CeO2 catalyst electrode in electrochemical cell. Appl Catal A Gen 2002; https://doi.org/10.1016/S0926-860X(01)00735-9. König et al., 2019 König M, Vaes J, Klemm E, Pant D. Solvents and ing electrolytes in the electrocatalytic reduction of CO2. iScience 2019; https://doi.org/10.1016/j.isci.2019.07.014. Kortlever et al., 2015a Kortlever R, Peters I, Koper S, Koper MTM. Electrochemical CO2 reduction to formic acid at low overpotential and with high Faradaic efficiency on carbon-ed bimetallic Pd-Pt nanoparticles. ACS Catal 2015a; https://doi.org/10.1021/acscatal.5b00602. Kortlever et al., 2015b Kortlever R, Shen J, Schouten KJP, Calle-Vallejo F, Koper MTM. Catalysts and reaction pathways for the electrochemical reduction of carbon dioxide. J Phys Chem Lett. 2015b;6:4073–4082 https://doi.org/10.1021/acs.jpclett.5b01559. Kuhl et al., 2012 Kuhl KP, Cave ER, Abram DN, Jaramillo TF. New insights into the electrochemical reduction of carbon dioxide on metallic copper surfaces. Energy Environ Sci 2012; https://doi.org/10.1039/c2ee21234j.

Lee et al., 2018 Lee SY, Jung H, Kim NK, Oh HS, Min BK, Hwang YJ. Mixed copper states in anodized Cu electrocatalyst for stable and selective ethylene production from CO2 reduction. J Am Chem Soc 2018; https://doi.org/10.1021/jacs.8b02173. Lee et al., 2019 Lee J, Lim J, Roh CW, Whang HS, Lee H. Electrochemical CO2 reduction using alkaline membrane electrode assembly on various metal electrodes. J CO2 Util 2019; https://doi.org/10.1016/j.jcou.2019.03.022. Liang et al., 2020 Liang S, Altaf N, Huang L, Gao Y, Wang Q. Electrolytic cell design for electrochemical CO2 reduction. J CO2 Util. 2020;35:90–105 https://doi.org/10.1016/j.jcou.2019.09.007. Liu et al., 2001 Liu M, He P, Luo JL, Sanger AR, Chuang KT. Performance of a solid oxide fuel cell utilizing hydrogen sulfide as fuel. J Power Sources. 2001;94:20–25 https://doi.org/10.1016/S0378-7753(00)00660-1. Logan et al., 2008 Logan BE, Call D, Cheng S, et al. Microbial electrolysis cells for high yield hydrogen gas production from organic matter. Environ Sci Technol 2008; https://doi.org/10.1021/es801553z. Lorenzi et al., 2017 Lorenzi G, Lanzini A, Santarelli M, Martin A. Exergoeconomic analysis of a direct biogas upgrading process to synthetic natural gas via integrated high-temperature electrolysis and methanation. Energy 2017; https://doi.org/10.1016/j.energy.2017.11.080. Lu et al., 2014 Lu Q, Rosen J, Zhou Y, et al. A selective and efficient electrocatalyst for carbon dioxide reduction. Nat Commun 2014; https://doi.org/10.1038/ncomms4242. Ma et al., 2016 Ma S, Sadakiyo M, Luo R, Heima M, Yamauchi M, Kenis PJA. One-step electrosynthesis of ethylene and ethanol from CO2 in an alkaline electrolyzer. J Power Sources 2016; https://doi.org/10.1016/j.jpowsour.2015.09.124. Maegaard et al., 2019 Maegaard K, Garcia-Robledo E, Kofoed MVW, et al. Biogas upgrading with hydrogenotrophic methanogenic biofilms. Bioresour Technol 2019; https://doi.org/10.1016/j.biortech.2019.121422. Manthiram et al., 2014 Manthiram K, Beberwyck BJ, Alivisatos AP. Enhanced

electrochemical methanation of carbon dioxide with a dispersible nanoscale copper catalyst. J Am Chem Soc 2014; https://doi.org/10.1021/ja5065284. Mao, 1991 Mao Z. A modified electrochemical process for the decomposition of hydrogen sulfide in an aqueous alkaline solution. J Electrochem Soc. 1991;138:1299 https://doi.org/10.1149/1.2085775. Mbah et al., 2010 Mbah J, Srinivasan S, Krakow B, et al. Effect of RuO2-CoS2 anode nanostructured on performance of H2S electrolytic splitting system. Int J Hydrogen Energy. 2010;35:10094–10101 https://doi.org/10.1016/j.ijhydene.2010.08.023. Mejia Likosova et al., 2014 Mejia Likosova E, Keller J, Poussade Y, Freguia S. A novel electrochemical process for the recovery and recycling of ferric chloride from precipitation sludge. Water Res. 2014;51:96–103 https://doi.org/10.1016/j.watres.2013.12.020. Merino-Garcia et al., 2018 Merino-Garcia I, Albo J, Irabien A. Tailoring gasphase CO2 electroreduction selectivity to hydrocarbons at Cu nanoparticles. Nanotechnology 2018; https://doi.org/10.1088/1361-6528/aa994e. Merino-Garcia et al., 2019 Merino-Garcia I, Albo J, Solla-Gullón J, Montiel V, Irabien A. Cu oxide/ZnO-based surfaces for a selective ethylene production from gas-phase CO2 electroconversion. J CO2 Util 2019; https://doi.org/10.1016/j.jcou.2019.03.002. Na et al., 2019 Na J, Seo B, Kim J, et al. General technoeconomic analysis for electrochemical coproduction coupling carbon dioxide reduction with organic oxidation. Nat Commun. 2019;10:5193 https://doi.org/10.1038/s41467-01912744-y. Narayanan et al., 2011 Narayanan SR, Haines B, Soler J, Valdez TI. Electrochemical conversion of carbon dioxide to formate in alkaline polymer electrolyte membrane cells. J Electrochem Soc 2011; https://doi.org/10.1149/1.3526312. Narendranath et al., 2019 Narendranath J, Manokaran J, Shanmuga Sundar S, et al. Electrochemical recovery of hydrogen and elemental sulfur from hydrogen sulfide gas by two-cell system. Energy Sources A Recover Util Environ Eff. 2019;00:1–14 https://doi.org/10.1080/15567036.2019.1680766.

Nelabhotla and Dinamarca, 2019 Nelabhotla ABT, Dinamarca C. Bioelectrochemical CO2 reduction to methane: MES integration in biogas production processes. Appl Sci. 2019;9:16–18 https://doi.org/10.3390/app9061056. Nitopi et al., 2019 Nitopi S, Bertheussen E, Scott SB, et al. Progress and perspectives of electrochemical CO2 reduction on copper in aqueous electrolyte. Chem Rev. 2019;119:7610–7672 https://doi.org/10.1021/acs.chemrev.8b00705. Ntagia et al., 2020 Ntagia E, Fiset E, Truong Cong Hong L, Vaiopoulou E, Rabaey K. Electrochemical treatment of industrial sulfidic spent caustic streams for sulfide removal and caustic recovery. J Hazard Mater. 2020;388:121770 https://doi.org/10.1016/j.jhazmat.2019.121770. Pan et al., 2019 Pan F, Li B, Deng W, et al. Promoting electrocatalytic CO2 reduction on nitrogen-doped carbon with sulfur addition. Appl Catal B Environ 2019; https://doi.org/10.1016/j.apcatb.2019.04.025. Papurello and Lanzini, 2018 Papurello D, Lanzini A. SOFC single cells fed by biogas: experimental tests with trace contaminants. Waste Manage 2018; https://doi.org/10.1016/j.wasman.2017.11.030. Park et al., 2019 Park S, Kim Y, Noh Y, et al. A sulfur-tolerant cathode catalyst fabricated with in situ exsolved CoNi alloy nanoparticles anchored on a Ruddlesden-Popper for CO2 electrolysis. J Mater Chem A 2019; https://doi.org/10.1039/c9ta07700f. Pasternak et al., 2011 Pasternak G, Rutkowski P, Śliwka E, Kołwzan B, Rybak J. Broad coal tar biodegradative potential of Rhodococcus erythropolis B10 strain isolated from former gasworks site. Water Air Soil Pollut. 2011;214:599–608 https://doi.org/10.1007/s11270-010-0449-2. Pasternak et al., 2019 Pasternak G, Yang Y, Santos BB, Brunello F, Hanczyc MM, Motta A. Regenerated silk fibroin membranes as separators for transparent microbial fuel cells. Bioelectrochemistry. 2019;126:146–155 https://doi.org/10.1016/j.bioelechem.2018.12.004. Pasternak et al., 2020 Pasternak G, Askitosari TD, Rosenbaum MA. Biosurfactants and synthetic surfactants in bioelectrochemical systems: a minireview. Front Microbiol. 2020;11 https://doi.org/10.3389/fmicb.2020.00358.

Pawar et al., 2019 Pawar AU, Kim CW, Nguyen-Le MT, Kang YS. General review on the components and parameters of photoelectrochemical system for CO2 reduction with in situ analysis. ACS Sustain Chem Eng 2019; https://doi.org/10.1021/acssuschemeng.8b06303. Persson, 2003 Persson M. Evaluation of upgrading techniques for biogas. Environ Eng. 2003; SGC142. Pikaar et al., 2015 Pikaar I, Likosova EM, Freguia S, Keller J, Rabaey K, Yuan Z. Electrochemical abatement of hydrogen sulfide from waste streams. Crit Rev Environ Sci Technol. 2015;45:1555–1578 https://doi.org/10.1080/10643389.2014.966419. Qiao et al., 2018 Qiao L, Bai J, Luo T, et al. High yield of H2O2 and efficient S recovery from toxic H2S splitting through a self-driven photoelectrocatalytic system with a microporous GDE cathode. Appl Catal B Environ. 2018;238:491– 497 https://doi.org/10.1016/j.apcatb.2018.07.051. Ren et al., 2019 Ren S, Joulié D, Salvatore D, et al. Molecular electrocatalysts can mediate fast, selective CO2 reduction in a flow cell. Science 2019; https://doi.org/10.1126/science.aax4608. Ritchie and Roser, 2020 Ritchie H, Roser M. CO2 and Greenhouse Gas Emissions Our World Data 2020. Rosen et al., 2011 Rosen BA, Salehi-Khojin A, Thorson MR, et al. Ionic liquidmediated selective conversion of CO2 to CO at low overpotentials. Science 2011; https://doi.org/10.1126/science.1209786. Rousseau et al., 2020 Rousseau R, Etcheverry L, Roubaud E, Basséguy R, Délia ML, Bergel A. Microbial electrolysis cell (MEC): strengths, weaknesses and research needs from electrochemical engineering standpoint. Appl Energy 2020; https://doi.org/10.1016/j.apenergy.2019.113938. Ryckebosch et al., 2011 Ryckebosch E, Drouillon M, Vervaeren H. Techniques for transformation of biogas to biomethane. Biomass Bioenergy 2011; https://doi.org/10.1016/j.biombioe.2011.02.033. Salvatore et al., 2018 Salvatore DA, Weekes DM, He J, et al. Electrolysis of gaseous CO2 to CO in a flow cell with a bipolar membrane. ACS Energy Lett

2018; https://doi.org/10.1021/acsenergylett.7b01017. Sánchez et al., 2019 Sánchez OG, Birdja YY, Bulut M, Vaes J, Breugelmans T, Pant D. Recent advances in industrial CO2 electroreduction. Curr Opin Green Sustain Chem. 2019;16:47–56 https://doi.org/10.1016/j.cogsc.2019.01.005. Selvaraj et al., 2016 Selvaraj H, Chandrasekaran K, Gopalkrishnan R. Recovery of solid sulfur from hydrogen sulfide gas by an electrochemical membrane cell. RSC Adv. 2016;6:3735–3741 https://doi.org/10.1039/c5ra19116e. Shiratori et al., 2008 Shiratori Y, Oshima T, Sasaki K. Feasibility of directbiogas SOFC. Int J Hydrogen Energy 2008; https://doi.org/10.1016/j.ijhydene.2008.07.101. Shiratori et al., 2010 Shiratori Y, Ijichi T, Oshima T, Sasaki K. Internal reforming SOFC running on biogas. Int J Hydrogen Energy 2010; https://doi.org/10.1016/j.ijhydene.2010.05.064. Shri Prakash et al., 2014 Shri Prakash B, Senthil Kumar S, Aruna ST. Properties and development of Ni/YSZ as an anode material in solid oxide fuel cell: a review. Renew Sustain Energy Rev 2014; https://doi.org/10.1016/j.rser.2014.04.043. Spinner et al., 2012 Spinner NS, Vega JA, Mustain WE. Recent progress in the electrochemical conversion and utilization of CO2. Catal Sci Technol. 2012;2:19–28 https://doi.org/10.1039/c1cy00314c. Sun et al., 2018 Sun S, Awadallah O, Cheng Z. Poisoning of Ni-Based anode for proton conducting SOFC by H2S, CO2, and H2O as fuel contaminants. J Power Sources 2018; https://doi.org/10.1016/j.jpowsour.2017.12.056. Suryanto et al., 2019 Suryanto BHR, Du HL, Wang D, Chen J, Simonov AN, MacFarlane DR. Challenges and prospects in the catalysis of electroreduction of nitrogen to ammonia. Nat Catal 2019; https://doi.org/10.1038/s41929-019-02524. Szynkarczuk et al., 1994 Szynkarczuk J, Komorowski PG, Donini JC. Redox reactions of hydrosulphide ions on the platinum electrode-I The presence of intermediate polysulphide ions and sulphur layers. Electrochim Acta 1994; https://doi.org/10.1016/0013-4686(94)EC075-I.

Verbeeck et al., 2019 Verbeeck K, De Vrieze J, Biesemans M, Rabaey K. Membrane electrolysis-assisted CO2 and H2S extraction as innovative pretreatment method for biological biogas upgrading. Chem Eng J. 2019;361:1479–1486 https://doi.org/10.1016/j.cej.2018.09.120. Wang et al., 2017 Wang Y, Liu T, Lei L, Chen F. High temperature solid oxide H2O/CO2 co-electrolysis for syngas production. Fuel Process Technol 2017; https://doi.org/10.1016/j.fuproc.2016.08.009. Wang et al., 2018 Wang G, Pan J, Jiang SP, Yang H. Gas phase electrochemical conversion of humidified CO2 to CO and H2 on proton-exchange and alkaline anion-exchange membrane fuel cell reactors. J CO2 Util 2018; https://doi.org/10.1016/j.jcou.2017.11.010. Wang et al., 2019 Wang Y, Lin H, Hu B. Electrochemical removal of hydrogen sulfide from swine manure. Chem Eng J. 2019;356:210–218 https://doi.org/10.1016/j.cej.2018.08.171. Whang et al., 2019 Whang HS, Lim J, Choi MS, Lee J, Lee H. Heterogeneous catalysts for catalytic CO2 conversion into value-added chemicals, BMC. Chem Eng. 2019;1:1–19 https://doi.org/10.1186/s42480-019-0007-7. Whipple et al., 2010 Whipple DT, Finke EC, Kenis PJA. Microfluidic reactor for the electrochemical reduction of carbon dioxide: the effect of pH. Electrochem Solid State Lett 2010; https://doi.org/10.1149/1.3456590. Wiener et al., 2006 Wiener MS, Salas BV, Quintero-Núñez M, Zlatev R. Effect of H2S on corrosion in polluted waters: a review. Corros Eng Sci Technol 2006; https://doi.org/10.1179/174327806X132204. Wilhelm et al., 2001 Wilhelm DJ, Simbeck DR, Karp AD, Dickenson RL. Syngas production for gas-to-liquids applications: technologies, issues and outlook. Fuel Process Technol 2001; https://doi.org/10.1016/S03783820(01)00140-0. Xie et al., 2011 Xie K, Zhang Y, Meng G, Irvine JTS. Electrochemical reduction of CO2 in a proton conducting solid oxide electrolyser. J Mater Chem. 2011;21:195–198 https://doi.org/10.1039/c0jm02205e. Xu and Carter, 2019 Xu S, Carter EA. Theoretical insights into heterogeneous

(photo)electrochemical CO2 reduction. Chem Rev 2019; https://doi.org/10.1021/acs.chemrev.8b00481. Xu et al., 2010 Xu C, Zondlo JW, Gong M, Elizalde-Blancas F, Liu X, Celik IB. Tolerance tests of H2S-laden biogas fuel on solid oxide fuel cells. J Power Sources. 2010;195:4583–4592 https://doi.org/10.1016/j.jpowsour.2010.02.078. Xu et al., 2016 Xu T, Scafa N, Xu LP, et al. Electrochemical hydrogen sulfide biosensors. Analyst. 2016;141:1185–1195 https://doi.org/10.1039/c5an02208h. Xu et al., 2020 Xu Y, Edwards JP, Zhong J, et al. Oxygen-tolerant electroproduction of C2 products from simulated flue gas. Energy Environ Sci. 2020;13:554–561 https://doi.org/10.1039/c9ee03077h. Xue et al., 2019 Xue X, Chen R, Yan C, et al. Review on photocatalytic and electrocatalytic artificial nitrogen fixation for ammonia synthesis at mild conditions: advances, challenges and perspectives. Nano Res 2019; https://doi.org/10.1007/s12274-018-2268-5. Yang et al., 2018 Yang HB, Hung SF, Liu S, et al. Atomically dispersed Ni(i) as the active site for electrochemical CO2 reduction. Nat Energy 2018; https://doi.org/10.1038/s41560-017-0078-8. Yoo et al., 2013 Yoo M, Han SJ, Wee JH. Carbon dioxide capture capacity of sodium hydroxide aqueous solution. J Environ Manage 2013; https://doi.org/10.1016/j.jenvman.2012.10.061. Yu et al., 2018 Yu L, Wang J, Hu X, Ye Z, Buckley C, Dong D. A nanocatalyst network for electrochemical reduction of CO2 over microchanneled solid oxide electrolysis cells. Electrochem Commun 2018; https://doi.org/10.1016/j.elecom.2017.11.019. Zaman and Chakma, 1995 Zaman J, Chakma A. Production of hydrogen and sulfur from hydrogen sulfide. Fuel Process Technol. 1995;41:159–198 https://doi.org/10.1016/0378-3820(94)00085-8. Zhao and Wang, 2016 Zhao C, Wang J. Electrochemical reduction of CO2 to formate in aqueous solution using electro-deposited Sn catalysts. Chem Eng J. 2016;293:161–170 https://doi.org/10.1016/j.cej.2016.02.084.

Chapter 10

Siloxanes removal from biogas and emerging biological techniques

Kazimierz Gaj, Department of Environment Protection Engineering, Wroclaw University of Science and Technology, Wrocław, Poland

Abstract

This chapter critically reviews and discusses the state-of-the-art technologies for the removal of siloxanes from biogas, indicating potentially beneficial development directions and deficiencies in the state of knowledge. The origin of volatile organic silicon compounds (VOSCs) in biogas, their selected physicochemical properties, and technological problems that they can cause during the energetic use of biogas are presented. Both implemented and at the experimental stage, methods of VOSC removal from landfill and sewage gases are reviewed and systematized. Mechanisms and parameters of purification processes, their effectiveness and efficiency are discussed, and individual methods are compared. Possibilities of regeneration of the spent sorbents are assessed. Premises and perspectives for their application are given. Finally, technologies worth further development, combining selected methods with each other and prospective directions for further research, are suggested.

Keywords

Biogas; volatile organic silicon compounds; volatile methylsiloxanes; adsorption; absorption; cooling; condensation; membrane separation; biological treatment

Chapter outline

Outline


10.2 Methods for reducing the content of volatile organic silicon compounds in biogas 258

10.2.1 Pretreatment methods 258 10.2.2 Refrigeration and freezing methods 259 10.2.3 Adsorption methods 260 10.2.4 Absorption methods 276 10.2.5 Membrane techniques 277 10.2.6 Biological methods 278

10.3 Combined methods for volatile organic silicon compounds removal from biogas 281

10.4 Comparison of the methods for reducing the content of volatile organic silicon compounds in biogas 281

10.5 Conclusions and future perspective 286

References 288

10.1 Introduction

Siloxanes are synthetic volatile oligomers, containing alternately connected silicon and oxygen atoms, with Si atoms being associated with hydrocarbon functional groups, usually methyl. Due to their volatility under typical ambient conditions, they are referred to as volatile methylsiloxanes (VMSs). Biogas may also contain intermediate siloxane degradation products—silanes and silanols, which, together with VMSs, have been designated as volatile organic silicon compounds (VOSCs) (Table 10.1).

Table 10.1

Chemical name

Abbreviation

Chemical formula

Vapor pressure, kPa, at 25°C

Hexamethylcyclotrisiloxane

D3

Si3-O3-(CH3)6

1.16

Octamethylcyclotetrasiloxane

D4

Si4-O4-(CH3)8

0.14

Decamethylcyclopentasiloxane

D5

Si5-O5-(CH3)10

0.03

Dodecamethylcyclohexasiloxane

D6

Si6-O6-(CH3)12

0.003

Hexamethyldisiloxane

L2

Si2-O-(CH3)6

5.613

Octamethyltrisiloxane

L3

Si3-O2-(CH3)8

0.445

Decamethyltetrasiloxane

L4

Si4-O3-(CH3)10

0.05

Dodecamethylpentasiloxane

L5

Si5-O4-(CH3)12

0.013

Trimethylsilanol

TMS-OH

Si-(CH3)3-OH

1.84

Tetramethylsilane

TMS

Si-(CH3)4

95.03

Notes: Structural patterns and other properties of VOSCs can be found, for example, in Gaj (2018). VOSCs, volatile organic silicon compounds.

VOSCs are the basic components of silicone polymers, the use of which in various branches of the economy and in personal care products has increased rapidly in recent years. They owe this to their unique properties, especially their high thermal stability, hydrophobicity, high resistance to oxidation and UV radiation, low surface tension and viscosity, high compressibility, low chemical reactivity, low crystallization temperature, and high spreadability. Most of them are colorless and odorless fluids, well-permeable to UV radiation and visible light. These features particularly predispose VMSs for use in such personal hygiene products as shampoos, hair sprays, creams, soaps, antiperspirants, shaving foams, etc. Used cosmetics, through sewage systems, flow to sewage treatment plants (STPs), and the packaging ends in landfills. From there, due to the elevated temperature intensifying desorption, VOSCs contained in these products volatilize into biogas. Siloxanes D4 and D5 predominate in biogas from sewage sludge. According to Kazuyuki et al. (2007), D5 constitutes over 90% of VOSCs contained in municipal sewage. They release into the air mainly during sewage aeration or adsorb into activated sludge flocs, with which they end up in the fermentation chambers. Those that are more soluble in water and with a lower vapor pressure, for example, L3, TMS, and TMS-OH, get into receivers of treated wastewater or into leachate in landfills. Therefore their content in biogas is usually insignificant. As a result of biogas combustion, VOSCs are oxidized, forming hard, difficult to remove deposits, consisting mainly of silica and silicates. These deposits cover the surface of combustion chambers, heat exchangers, cylinder heads, valves and spark plugs, fuel cell electrodes, nozzles, turbine blades, etc. In addition, they get into engine oils, deteriorating their lubricating properties. They are able also to deactivate catalysts used for exhaust gas treatment. This all results in uneven operation of the devices, their faster wear, and clogging of ducts and nozzles, which can lead to frequent and expensive servicing or complete seizure of an

engine. For these reasons, VOSCs are among the most troublesome trace compounds present in biogas, significantly affecting the efficiency of biogas plants. In many cases, to ensure compliance with warranty conditions and limit the aforementioned threats, it has become necessary to remove them from biogas. However, a serious obstacle in this field is the physicochemical diversity of this group of compounds, their poor water solubility, relatively poor chemical reactivity and resistance to physical interactions, as well as the special ability to adsorb on various materials, due to the relatively high organic carbon/water partition coefficients (KOC) and octanol/air partition coefficients (KOA), especially for D5 and D6 (Table 10.1). In addition to biogas treatment, which is expensive and technically difficult, attempts have been made to remove VOSCs and their precursors from sewage and sewage sludge. The permissible content of VOSCs in biogas, specified in technical and warranty conditions, significantly differs depending on the type of device using biogas and the manufacturer (Gaj, 2017; Zamorska-Wojdyła et al., 2012. The most resistant in this respect are traditional boilers and gas piston engines. In their case, economic analysis should decide whether it is more profitable to pay for more frequent engine maintenance or for biogas processing. Since the removal of VOSCs from biogas is technologically troublesome and they have a significant impact on the energy efficiency of a biogas plant, a pretreatment method and more frequent maintenance may sometimes be the more beneficial alternative. However, when using catalytic afterburners, fuel cells, turbines, or microturbines, removal of VOSCs seems to be unavoidable, since their concentrations in biogas usually exceed the manufacturer’s limits. It may also be required when introducing biogas into the natural gas grid. For economic reasons, it would be preferably to remove VOSCs as a part of a comprehensive biogas treatment system (Gaj, 2017).

10.2 Methods for reducing the content of volatile organic silicon compounds in biogas

The following conventional technologies are available for biogas purification from VOSCs: adsorption, absorption, and cooling. Emerging technologies that are still at a research phase include adsorption with unconventional adsorbents [e.g., some polymer adsorbents (PAs) and new zeolites (ZEs)], membrane techniques, and biological methods. Removal of VOSCs can be also carried out initially, before the anaerobic digestion process, ing any subsequent biogas treatment.

10.2.1 Pretreatment methods

The primary pretreatment method is desorption of VOSCs from sewage sludge, for example, during more intense aeration in the biological stage of STP or between sludge heating and digestion, with optional thermal treatment of sludge. For example, Oshita et al. (2014) showed that by subjecting prefermentation sludge to heat treatment and intensive aeration, approximately 90% of VMSs can be stripped off. The tendency to distribute VOSCs from water to air is due to the high Henry constants and high octanol–water partition coefficients (Gaj, 2018). In addition, the VMS removal efficiency can be increased by adding sodium hydroxide to the pretreatment process (Oshita et al., 2015). Thermal treatment may be economically advantageous, especially in the case of thermophilic fermentation. Chemical oxidation, for example, by peroxymonosulfate and dimethyldioxirane (Appels et al., 2008), is another way to reduce the VOSCs content in sewage sludge before fermentation. This method breaks down larger siloxane oligomers into smaller ones and creates mineral deposits that are easy to separate in the sedimentation stage. It also causes the decomposition of extracellular polymeric

substances, forming a biofilm covering the surfaces of devices immersed in wastewater on which VOSCs are adsorbed. As a result of the above mechanisms, some VOSCs are released into the atmosphere or precipitate in the form of silicates or silica. An additional advantage of these processes is increasing the intensity of the CH4 production in a fermentation chamber by reducing the time needed for the hydrolysis and limiting fermentation inhibitors, such as NH3 and compounds of sulfur and chlorine. Thermal disintegration of microorganism cells also facilitates sludge dewatering. On the other hand, the intensification of sewage sludge aeration and heating may pose a threat to air quality (Gaj and Pakuluk, 2015), which should be prevented, for example, by biofiltration of stripped gases. Electrochemical oxidation of VOSCs with hydroxyl radicals is also possible, and can be used, for example, for simultaneous removal of persistent organic pollutants from wastewater (Zhang et al., 2016). In contrast to the mechanisms described above, biodegradation of VMSs during the STP process is not significant (de Arespacochaga et al., 2015).

10.2.2 Refrigeration and freezing methods

Cooling/condensation and freezing techniques allow the simultaneous removal of moisture, some VOSCs, and other volatile compounds from biogas. These techniques can be advantageous before the adsorption process, ensuring longer adsorbent lifetime, especially when using silica gel (SG) or activated carbon (AC). Laboratory tests for removing VOSCs (L2, D3, D4, D5, TMS-OH) in a refrigeration condenser were conducted by Schweigkofler and Niessner (2001). In biogas dried this way to a dew point of 5°C they found ~10% reduction in VOSCs content. In turn, full-scale research by Wheless and Gary (2002) showed the possibility of 50% total VMSs removal as a result of cooling biogas to about 5°C, at a pressure of 2.5 MPa. However, they are removed with varying efficacy, according to their saturation partial pressure and biogas humidity. Therefore this value cannot be reliable for any biogas. Generally, the VOSCs removing efficacy of the refrigeration/condensation techniques is typically from a few to 50%, depending on the cooling temperature, gas pressure, gas humidity, and

composition of the VOSCs mixture. However, the removal mechanism is not sufficiently recognized. It can be suspected that at a temperature just below the water dew point, the VOSCs removal effect is more a result of their washout by the condensate than direct condensation. The latter mechanism may be relevant when biogas is cooled more intensely, to a temperature of about −30°C. In these conditions the removal of VMSs can reach 95% (Wheless and Pierce, 2004). Generally, cooling to approximately 5°C, at atmospheric pressure, as is usually practiced in biogas drying processes, allows only slight removal of VOSCs. It is particularly difficult to remove more volatile VMSs (L2, D3, L3) this way. Additionally, cooling below 0°C can cause freezing and blocking of the plant. In this case, periodic defrosting is required. The cryogenic cooling, possibly combined with biogas compression, due to high energy consumption, does not seem to be economically justified. It can be profitable only for high concentrations of VOSCs (~several hundred of mg/m³) or in the case of simultaneous CO2 removal (CO2 boiling point at 0.1 MPa is −78.5°C), or the production of liquefied biomethane (CH4 boiling point at 0.1 MPa is −161.5°C). In the latter case, however, there is a risk of precipitating CH4 clathrates, especially under elevated pressure, and blocking the installation.

10.2.3 Adsorption methods

Physical or physicochemical densification of gas impurities on an active, expanded surface of various, both organic and mineral, porous solids is one of the basic techniques for gas purification. A specific surface area, total pore volume, and sizes of meso- and micropores of an adsorbent are decisive for the process effectiveness. The latter are crucial and should be slightly larger than the size of molecules removed. In the case of biogas purification, adsorbents such as AC, SG, ZEs, and alumina have found commercial use. Emerging adsorbents include polymeric resins, on which numerous studies are currently underway. For adsorption of VOSCs it is also possible to adapt active oxides of other metals besides Al, minerals such as kaolinite, illite, mica, hematite, diatomite, Fuller earth, meadow ore, and others, which, due to space limitations, will be omitted.

10.2.3.1 Adsorption into activated carbon

AC is a substance consisting mainly of elemental carbon in an amorphous form, characterized by the especially expanded specific surface area (Table 10.2). This largest unit surface area among all adsorbents results from a very developed porous structure, consisting mainly of micropores (<2 nm) and mesopores (2– 50 nm). AC is obtained from organic raw materials (fossil coals, petroleum coke, peat, wood, and some waste products of the agrifood industry) in the processes of carbonization and activation. Carbonization is based on thermal treatment (600°C–800°C), during which organic matter is decomposed and volatile parts are removed, whereby carbonizate obtains a porous structure. Activation can be carried out with steam, CO2, and O2 at a high temperature (at 400°C–500°C—to eliminate the rest of the volatile matter, at 800°C–1000°C—to develop the porosity by partial gasification), or by roasting (at 500°C–900°C) the raw material with such chemicals as zinc chloride, phosphoric acid, potassium sulfide, or potassium thiocyanate (Yang, 2003). In general, the VOSCs adsorption is more effective into carbons activated with H3PO4 (CabreraCodony et al., 2018). Activation with H3PO4 provides a larger pore volume, which facilitates the capture of the larger siloxane molecules, that is, D4 and D5. However, smaller and more volatile VMSs, like the L2, are easily desorbed after the breakthrough.

Table 10.2

Parameter Specific surface area (BET)

Number of ACs tested 6

Value 650–1370

Unit m²/g

12

850–1757

Cabrera-Codony et al. (2014)

5a

909–1250


5b

1082–2190

11

262–1573

Yu et al. (2013)

3

302c; 879; 1104

Nam et al. (2013)

12

819–1323

Vagenknechtová et al. (2017)

1

1220

Ortega and Subrenat (2009)

1

930

Sigot et al. (2016)

Average pore diameter

6

2.01–3.04

3

1.65; 2.34

Nam et al. (2013)

1

0.6–3.0

Sigot et al. (2016)

Total pore volume

6

0.35–0.94

3

0.45; 0.47; 0.18c

Nam et al. (2013)

12

0.39–1.38


1

0.67


12

0.45–1.52


5a

0.42–1.10


5b

0.91–1.58

nm

cm³/g

AC, activated carbon; VMSs, volatile methylsiloxane.

aSteam activation.

bH3PO4 activation.

c2.5% KOH impregnation.

The tests confirmed the clear effect of the Brunauer-Emmett-Teller (BET) surface area and the volume of micropores (1–2 nm) and small mesopores (2– 6 nm) on the VMS adsorption capacity. It is worth emphasizing that impregnation with 2.5% KOH significantly reduces the pore volume and the specific surface area of the tested ACs. The important feature that distinguishes AC from other major adsorbents is its nonpolar (or only slightly polar) nature (Soreanu et al., 2011; Yang, 2003), which promotes the adsorption of nonpolar and weakly polar molecules, such as VOSCs. ACs, in addition to pure carbon, contain various impurities that affect their adsorption capacity. They are both mineral substances (mainly ash) and single heteroatoms (O, H, S, Cl). Their share depends on the type of raw material used and the activation method. Mineral substances are not chemically bound. They are removed by leaching with concentrated hydrochloric or hydrofluoric acid. In turn, single heteroatoms are chemically bonded to a carbon skeleton (Repelewicz and Choma, 2006). They give the adsorbent an acidic or basic character. Matsui and Imamura (2010) and Gong et al. (2015) indicated that the AC activity toward VMSs increases with increasing pH. On the other hand, alkaline impregnation can reduce the AC active surface and the pore volume (Table 10.2). Finocchio et al. (2009) showed that alkali-impregnated AC have a lower D3 adsorption capacity than pure AC. The authors also tested impregnation with

Cr(VI) and Cu(II) salts, showing that it works favorably for D3 adsorption. However, this method creates environmental threats. For high concentrations of H2S, Wheless and Pierce (2004) proposed the use of two-stage adsorption. The first, adsorber with AC impregnated by potassium permanganate or sodium hydroxide, for the preliminary removal of H2S, and the second, with pure AC, for the final removal of H2S and VOSCs. This can significantly extend the lifetime of the adsorbent at the second stage and reduce total purification costs.

Adsorption capacity of activated carbon

AC has highly diverse micro- and mesopores and a large fraction of pores with a diameter corresponding to the diameter of the VOSC molecules. However, there are not only VOSCs. Its structure also allows adsorption of other undesirable biogas components, for example, sulfur compounds, water vapor, halides, and other volatile organic compounds (VOCs). In some cases this feature can be an advantage, but it significantly reduces the adsorbent lifetime and reduces its activity toward VOSCs. Also, AC adsorbs nondisturbing ingredients, for example, hydrocarbons other than CH4 (Urban et al., 2009), as well as small amounts of valuable CH4. According to Yang (2003), CH4 adsorption into AC is two times greater than into a 5 Å molecular sieve and three times greater than into an SG. Additionally, the presence of alkanes and aromatics in biogas that compete for a place in the AC micropores reduces its adsorption capacity relative to VOSCs. AC adsorption capacity is different for various siloxanes. Lighter chain molecules (e.g., L2) break the bed faster than the heavier and larger cyclic molecules (e.g., D5). Wheless and Pierce (2004) found that biogas without L2, or with only a low content, cleans itself more effectively than biogas with a high content of this compound. Their research, using real, dried (dew point ~5°C) and compressed (~2.5 MPa) biogas, showed that AC under these conditions can adsorb from 10 to 15 g VMSs/kg AC. A similar order of VMS adsorption capacity (8.5–21.5 g/kg) also was obtained by Hepburn et al. (2015). The results of adsorption capacity tests for various ACs and biogases are summarized in Table 10.3.

Table 10.3

Kind of tested biogas

Kind of adsorbate

Adsorption capacity, g/kg

Real sewage biogas

Total VMSs

8.5–21.5

5

de Arespacochaga et al. (2014)

10–15

Wheless and Pierce (2004)

D4 + D5

63


Synthetic biogas (CH4/CO2=50/50, vol/vol)

D4

43–334

22a–225b

Yu et al. (2013)

Synthetic biogas (carrier gas: air)

L2

110–350

Synthetic biogas (carrier gas: N2)

10–100

Gislon et al. (2013)

D4

53

Sigot et al. (2016)

56–192

Matsui and Imamura (2010)

50–404

Oshita et al. (2010)

D5

50–531

106

Kajolinna et al. (2015)

D6

18

Total VMSs

155–307

L2

47–123

D4

36–90

D5

47–169

Nam et al. (2013)

AC, activated carbon; VMSs, volatile methylsiloxane.

aBET: 432 m²/g, micropore volume (Vmic): 0.18 cm³/g.

bBET: 1573 m²/g, Vmic=0.62 cm³/g.

The spread of the results shown in Table 10.3 is very wide. The highest AC adsorption capacities were usually found for cyclic VMSs. This may be related to their physical structure and their changes on AC surface, described below. It should be noted that the high results obtained for artificial biogas are not representative of real conditions (dry gas was usually used, with no competing pollutants).

Mutual displacement of volatile methylsiloxanes from activated carbon

Many researchers note the phenomenon of competing adsorption, that is, displacement of previously adsorbed lighter VMSs (e.g., L2 and L3) by heavier and less volatile ones (e.g., D4 and D5), or by such compounds as water vapor or high-molecular-weight, aromatic VOCs (Arnold and Kajolinna, 2010; Ajhar et al., 2010; Matsui and Imamura, 2010). This happens when AC becomes saturated. L2 is usually the first siloxane to break through, with its outlet concentration quickly reaching a higher value than the value at the inlet. The same happens with D4 and D5. Matsui and Imamura (2010) analyzed their breakthrough and found that D4 appears in outlet gas earlier than D5, and reaches a concentration higher than that at the inlet. On this basis, they hypothesized that D5 displaces previously adsorbed D4, which leads to an increase in D4 concentration in the outlet gas. The AC adsorption capacity and

breakthrough time are therefore strictly dependent on the shares of the individual VMSs in biogas. This phenomenon can be reduced by using an adsorbent with a larger proportion of micropores or using a mixture of adsorbents with more different pore sizes (de Arespacochaga et al., 2014). Precooling of biogas is also beneficial.

Preparation of biogas for adsorption into activated carbon

An important disadvantage of AC is its nonselectivity, which significantly reduces the VOSC adsorption capacity. This effect can be reduced by biogas pretreatment consisting of condensation drying (cooling to 2°C–5°C) or washing with, for example, Selexol (Wheless and Gary, 2002). The purpose of these treatments is primarily to reduce the relative humidity below 50% and to remove, together with condensate, some of the competing impurities. It is also desirable to ensure gas temperature in the range of 10°C–30°C (below 10°C there may be a problem with maintaining the relative humidity<50%, and above 30°C the AC activity decreases and desorption may occur). At high concentrations of VOSCs—in the range of 200–1000 mg/m³—it may be profitable to use freezing (<−30°C) prior to adsorption, which—in addition to drying the biogas—can provide preliminary removal of up to 90% of VOSCs and other volatile compounds (Prabucki et al., 2001). On the other hand, removal of oxygen-containing VOCs, for example, acetates, alcohols, and ketones (often found in landfill gas), reduces the risk of dissolving adsorbed VOSCs and their reevaporation into the biogas.

Possibilities of activated carbon regeneration

As a standard, regeneration of spent AC may be carried out thermally, at a temperature of 100°C–300°C, for several dozen minutes, using steam, nitrogen, or air as a carrier gas. For larger molecules, such as cyclic VMSs, the use of steam is most effective (Miguel et al., 2002). However, the desorption of VOSCs

is not complete, and the bed must be replaced frequently. Thermal regeneration is also quite expensive due to the relatively high boiling points of VOSCs (up to 245°C for D6—Table 10.1) and problematic—due to the increased risk of selfignition and potential destruction of the porous AC structure. Additionally, stripped VOSCs commonly end up in the environment, posing a potential threat to ecosystems (Gaj, 2018). For example, the D5 thermal desorption degree of spent AC in the tests of Schweigkofler and Niessner (2001) was 74%–83% (250°C, 20 min). In the case of the lighter and more volatile L2, it exceeded 95%. In turn, Ortega and Subrenat (2009) and Gislon et al. (2013) obtained after regeneration 40% (at 90°C) and 70% (at 200°C) of the L2 initial capacity, respectively. A phenomenon of the limited desorption of cyclic VMSs on the example of D3 was investigated by Finocchio et al. (2009). They showed that D3 undergoes transformation during adsorption. Based on the chemical analysis of the spent AC, previously treated with CO2/CH4/steam mixture with D3, they hypothesized that part of D3 is transformed into a nonvolatile compound with a much higher molar mass—polydimethylsiloxane (PDMS), which is hardly desorbable. They also found an accumulation of silicon compounds in spent AC. Thus, during adsorption, cyclic VMSs change their form into a chain one, and undergo polymerization. This process deactivates AC by blocking active pores and limits its regeneration. This phenomenon has also been confirmed by Giraudet et al. (2014) in research on D4 adsorption and AC regeneration (at 145°C). Also, Cabrera-Codony et al. (2014) showed polymerization of D4 to higher cyclic VMSs by detecting D5, D6, and D7 in saturated AC, while D4 was the sole source of silicon. According to them, a key role in the polymerization process is played by the presence of oxygen functional groups (proton-donating groups), especially carboxylic and phenolic groups that catalyze the breaking of rings. As a result, with the assistance of water vapor, lineal silanediols are formed during the VMS hydrolysis process, which then polymerize mainly to PDMS (Cabrera-Codony et al., 2014; Sigot et al., 2015). This process is more intensive in the case of AC activated with phosphoric acid than with steam (Cabrera-Codony et al., 2015). There were also attempts at chemical regeneration of spent AC. This consisted of the oxidation of adsorbed VMSs to more water-soluble compounds (silanes and silanols)—using O3, H2O2, and/or iron salts (Fenton process). In these cases, however, there is a risk of SiO2 formation and deposition into the AC pores and a decrease in its original porosity (Cabrera-Codony et al., 2015, 2017). The results of chemical regeneration of AC exposed to D4 with O3 and H2O2 (Cabrera-Codony et al., 2015) are not satisfactory so far—less than 50%

efficiency was achieved. The authors speculate that the reasons for such low efficiency may be incomplete D4 oxidation. They obtained much better results when using iron compounds with hydrogen peroxide (up to 92%), which— according to the authors—may be caused by the intensification of catalytic oxidation. Due to the relatively high energy demand, especially in the case of higher cyclic VMSs, the risk of self-ignition and the risk of environmental pollution, thermal regeneration does not seem a recommendable technique for AC renewal. Further research should be conducted on the possibilities of the chemical regeneration of AC.

10.2.3.2 Adsorption into silica gel

SG is a type of inorganic xerogel, formed synthetically from sodium silicate by polymerization of silicic acid under the action of H2SO4 or HCl. The result of the reaction is hydrated silicon dioxide molecules that polymerize to an amorphous form of a gel with the molecular formula (SiO2)·nH2O, after which it is dried and activated by calcination. Characteristic feature of the SG is the 3D tetrahedral structure with silanol groups (Si-OH) attached to its surface. It is, apart from AC, the most popular adsorbent, characterized by high porosity, mechanical strength, high temperature resistance (up to 900°C), nonflammability, as well as chemical and biological inertness. Due to its high porosity and polar structure, SG has a special affinity for water, making it one of the most effective gas desiccant agents. Compared to AC, it has a smaller specific surface area (Table 10.4) and a larger pore size range (from 0.5 to 300 nm), with the majority of mesopores having a typical equivalent diameter of ~5 nm.

Table 10.4

Parameter

Value

Specific surface area (BET)

690

607


470

Jung et al. (2017)

717


656

Nam et al. (2013)

350

Montanari et al. (2010)


2.2

2.34

Nam et al. (2013)

Total pore volume

0.24

0.36

Nam et al. (2013)

Unit

Source

m²/g

Sigot et al. (2016)

nm

Soreanu et al. (2011)

cm³/g


SG, silica gel; VMS, volatile methylsiloxane.

Adsorption capacity and regeneration of silica gel

Good VMS adsorption properties of SGs (Table 10.5) have been confirmed in numerous tests (Schweigkofler and Niessner, 2001; Wheless and Pierce, 2004; EPRI, 2006; Montanari et al., 2010; Ortega and Subrenat, 2009; Sigot et al., 2014, 2015, 2016; Jung et al., 2017). For example, research on real biogas at the Calabasas landfill in California (US) showed that its adsorptive capacity relative to VMSs is about three times greater than with AC (EPRI, 2006). A higher SG adsorption activity was also confirmed by Wheless and Pierce (2004), Ryckebosch et al. (2011), and Schweigkofler and Niessner (2001). However, it is necessary to dry biogas to the relative humidity RH<10%. The latter authors also found that in contrast to AC, SG adsorbs L2 similarly or better than D5. Due to the higher affinity for L2 and lower affinity for sulfur compounds (Wheless and Pierce, 2004; EPRI, 2006), SG may therefore be a better adsorbent than AC for landfill gas treatment, which usually contains more L2 and less H2S than sewage gas.

Table 10.5

Kind of tested biogas Synthetic biogas, carrier gas: N2

RH, % 0

Kind of adsorbate D4

70

21

ND

104


0

D5

~100

25

77

50

~42

100

35

<10

L2, D5

>100

ND

Total VMSs

202

L2

17

D4

56

D5

129

91

Kajolinna et al. (2015)

D6

20

Synthetic biogas, carrier gas: air

70

L2

Synthetic biogas, CO2/CH4, 45/55, vol/vol

ND

D3

ND, No data; SG, silica gel; VMSs, volatile methylsiloxanes.

In general, most researchers indicate similar or slightly higher SG adsorption capacity relative to AC, provided that the gas is predried or the hydrophobicity of SG is increased by appropriate chemical impregnation. Otherwise, SG activity against VMSs drops rapidly (Table 10.5). The potential advantage of SG over AC may result from greater selectivity and potentially better susceptibility to regeneration of the former. However, literature reports on the thermal efficiency of SG regeneration are contradictory. According to Schweigkofler and Niessner (2001), in the case of L2 and D5, almost complete (>95%) desorption from SG can be obtained, while the degree of desorption of D5 from AC at the same conditions (250°C for 20 min) is only 74%–83%. On the other hand, Sigot et al. (2014) observed a 90% decrease in the activity of SG after desorption at 300°C [however, VMS load was almost 20 times higher than in the studies of Schweigkofler and Niessner (2001)]. Limited desorption of D3 from SG was also observed by Montanari et al. (2010). According to them, D3 is adsorbed into SG via a relatively stable hydrogen bond on the surface hydroxyl groups, which may promote breaking of the D3 bond and formation of silanols. Finally, D3 polymerizes to silicones, which—as in the case of AC—is the main reason for incomplete regeneration and its rapid deactivation. Thermal treatment up to 200°C causes only partial release of silanols groups. Sigot et al. (2016) also noted that H2S adsorption on SG drops sharply in the presence of D4, which may be due to the polymerization effect of the latter and blocking the pores. Of course, the susceptibility to desorption also depends on the volatility of stripped VOSCs, so it depends strictly on the composition of a particular biogas. Generally, due to the weaker adsorption forces of SG, thermal regeneration of spent SG should be easier than for AC and ZEs (Yang, 2003). An interesting adsorbent based on SG, which does not seem to have the above drawbacks, was developed by Liu et al. (2019). In their research on synthetic biogas (carrier gas—N2), acetylated hydrophobic SG was synthesized, via treatment of SG with the acetic anhydride. According to the authors, its

adsorption capacities reached 304 g/kg for L2 and 916 g/kg for D4, regardless of biogas humidity. The spent adsorbent could be easily regenerated by heating at 110°C (adsorption capacity of L2 and D4 was constant after four cycles).

10.2.3.3 Adsorption into zeolites

ZEs include hydrated aluminosilicates of metals from the first and second group of the periodic table (mainly Na, K, Ca), with a crystal structure formed on the basis of tetrahedrons (AlO4) and (SiO4) connected by oxygen atoms. As a result of thermal treatment, they lose water, obtaining a porous, unified spatial structure, which makes them selective adsorbents. ZEs are made of repetitive structural units forming channels and free spaces with a well-defined, narrow range of pore sizes. For this reason, they are also called molecular sieves. They can also be made synthetically. The tetrahedrons mentioned above are the fundamental elements of the two basic, commercially available synthetic ZEs, used in gas purification: A and X (Table 10.6). They are mechanically durable, resistant to acids and thermally stable (up to 600°C).

Table 10.6

Parameter

Kind of zeolite

Synthetic

Natural

3A (KA)

4A (NaA)

5A (CaA)

0.3

0.4

Nominal pore opening, nm

10X (CaX) 0.5

13X (NaX) 1.0

ZSM-5 1.0

Si/Al

0.7–1.2

1.0–1.5

5

4.5–5.0

1.6–3.0

The surface of high-silicon ZEs (Si/Al>10) is hydrophobic, and therefore—in contrast to low-silicon ZEs (Si/Al<1.5)—they are better suited for the adsorption of nonpolar and weakly polar substances, such as VOSCs (Table 10.7).

Table 10.7

Parameter

Type

Value


DAY40

607

ZE (cbv720)

780

Jung et al. (2017)

ZSM-5

419

Jiang et al. (2016)

UCT-15

424

13X

700

Sigot et al. (2016)

13X (MS544)

500

Montanari et al. (2010)

Hydrophobic ZE

712


Six synthetic ZEs

370–910



ND

3–9

Hydrophobic ZE

3


Unit

Source

m²/g

Ortega and Subrena

nm

Soreanu et al. (2011

Total pore volume

DAY40

0.33

ZSM-5

0.21

Jiang et al. (2016)

UCT-15

0.21

Hydrophobic ZE

0.51


cm³/g

Ortega and Subrena

ZEs, zeolites; VMS, volatile methylsiloxane.

Adsorption capacity and regeneration of zeolites

The D4 adsorption process using popular 13X molecular sieves has been studied, for example, in (Schweigkofler and Niessner, 2001), Japan (Matsui and Imamura, 2010), (Sigot et al., 2016), and Italy (Montanari et al., 2010). The results showed that they have a comparable or lower adsorption capacity for VMSs than AC and SG, and that L2 (similarly to AC and SG) quickly penetrates the bed. Matsui and Imamura (2010) compared the D4 adsorption capacity of 13X and 8A ZEs, showing the clear advantage of the former (Table 10.8), which—most likely—results from too small pores of the latter. In general, the use of ZEs to adsorb VMSs is problematic, because their pore sizes are usually smaller than the size of the adsorbate molecules. On the other hand, an interesting and unexplained so far phenomenon was observed— the relatively high degree of D4 adsorption on the ZSM-5 ZE, whose pores are generally smaller than the size of the D4 molecule. According to CabreraCodony et al. (2017), this effect may result from the catalytic conversion of cyclic VMSs into smaller, linear ones that can penetrate into the internal pores of ZEs.

Table 10.8

Kind of tested biogas

Kind of ZE

Kind of adsorbate

Synthetic biogas, carrier gas: N2

13X

D4

77


8A

4

13X (MS544)

D3

276

Hydrophobic ZE

D4

51

D5

52

ZE (cbv720)

19

Jung et al. (2017)

ZSM-5

D4

41

UCT-15

77

Clinoptilolite

11

Six synthetic ZEs

28–143

Synthetic biogas, carrier gas: air

DAY40


L2

ZEs, zeolites; VMS, volatile methylsiloxane.

An additional advantage of ZEs—in contrast to SGs—may be their good affinity for H2S, which is due to their alkaline nature. Moreover, they have greater hydrophobicity (especially high-silicon ZEs) compared to AC and SG. Ortega and Subrenat (2009) showed that the RH up to 70% does not affect the sorption capacity of DAY40 ZE. ZEs also show better thermal resistance, which allows their regeneration at temperatures above the boiling point of VMSs (preferably >250°C). Such high temperature can destroy the porous structure of AC. An important advantage of ZEs compared to AC is also the several times lower CH4 adsorption capacity (Yang, 2003). Regeneration of spent ZEs is carried out, as in the case of other adsorbents, by desorption at elevated temperature and/or under reduced pressure, using inert gas, or by displacing the adsorbate with a substance, which is then easily desorbed at elevated temperature. As in the case of AC, chemical regeneration is also possible. Currently, research is underway on new ZE adsorbents, among others in the United States (Jiang et al., 2016) and in Poland. The research in Poland concerns the possibility of using halloysite—a natural aluminosilicate mineral of volcanic origin, belonging to the group of kaolin minerals. The feature that makes halloysite particularly suitable for comprehensive biogas purification is its ability for simultaneous removal of VMSs, H2S, NH3, and NMVOCs (Gaj, 2017).

10.2.3.4 Adsorption into alumina

Alumina-based adsorbents are obtained by calcination of hydrated aluminum hydroxide (Al(OH)3·nH2O) in the presence of air, at a temperature of ~400°C.

As a result of dehydration, a crystalline, porous structure is obtained, consisting mainly of mesopores with a diameter of 3–7 nm and a specific surface area of 200–250 m²/g (Table 10.9). To obtain a larger surface area, on the order of 300– 400 m²/g, rapid roasting of the raw material at 400°C–800°C is carried out, as a result of which it acquires an amorphous form. Like SG, activated Al2O3 (AA) is polar, thus showing greater affinity for polar molecules (e.g., H2O, HF, NH3). It is therefore also a good desiccant. In addition to thermal treatment, the AA pore structure can be tailored by the action of acid (HCl or HF) or alkali (Yang, 2003).

Table 10.9

Parameter

Type

Value


Commercial AA

201

Synthetic Mesoporous AA (Al120-8h)

314

Commercial AA—Daejung (Korea)

250

Nam et al. (2013)


Commercial AA

3.4


4.3


6.7

Nam et al. (2013)

Total pore volume

Commercial AA

0.26


0.46


0.38

Nam et al. (2013)

Unit m²/g

nm

cm³/g

AA, activated Al2O3; VMS, volatile methylsiloxane.

Adsorption capacity and regeneration of activated Al2O3

According to Lee et al. (2001), AA exhibits a higher D4 adsorptivity compared to AC and ZE 13X. This has been confirmed by Nam et al. (2013) also in relation to D5. However, in the case of L2, the latter authors stated the opposite trend, which is probably the result of too large AA pores relative to the L2 molecule. In turn, Zhong et al. (2017) developed a synthetic adsorbent (Al1208h) based on alumina, demonstrating its greater D4 adsorption capacity compared to commercial AA (Table 10.10). More importantly, it turned out that it is easily and effectively regenerable (with capacity of 96%, 88%, and 85% after each subsequent cycle), without a significant D4 polymerization effect. Thus, they confirmed earlier reports by Lee et al. (2001), who obtained 90% regeneration of spent AA.

Table 10.10

Kind of tested medium

Kind of AA


Commercial AA


168

Adsorption equilibrium test by injecting of the VMS mixture into the AA


D4

34

D5

104

Total VMS

146

AA, activated Al2O3; VMS, volatile methylsiloxane.

10.2.3.5 Adsorption into polymer adsorbents

Synthetic polymer resins, known for use in chromatographic columns, have also been attempted for VOSC adsorption. They are obtained mainly by emulsion polymerization of monomers in the presence of an organic solvent. Most commercially available PAs are based on styrene cross-linked with divinylbenzene (DVB) or on acrylates (e.g., acrylic ester). The former include, for example, hydrophobic XAD-2 and XAD-4 resins, and the latter include hydrophilic XAD-7 resin (Yang, 2003). XAD polystyrene resins have a chemically homogeneous, nonionic structure. They are relatively durable and thermally stable (up to 200°C). Due to the dense cross-linking, they have a relatively large specific surface area and pores (Table 10.11), as well as high rigidity and mechanical strength. PAs can be easily modified by functional groups to achieve adsorption properties suitable for a particular adsorbate. Due to the aromatic surface formed mainly by benzene rings, PAs are usually highly hydrophobic, which particularly predisposes them—in contrast to ACs and SGs —to purify wet biogas.

Table 10.11

Parameter

Type

Value

Unit

Source


RS1

936

PDVB

831

Jafari et al. (2015)

PDVB-VI

594–780

P(DVB-ACAM)

271

Noshadi et al. (2016)

RPA

120

Jung et al. (2017)

XAD-2 (PS/DVB)

300

Yang (2003)

XAD-4 (PS/DVB)

725

XAD-7 (acrylic ester)

450


XAD-2 (PS/DVB)

XAD-4 (PS/DVB)

4

XAD-7 (acrylic ester)

9

PDVB

1.5–60


Total pore volume

PDVB

1.7

PDVB-VI

1.2–1.8

9

m²/g

nm

cm³/g

Oshita et al. (2010

PAs, polymer adsorbents; VMS, volatile methylsiloxane.

Adsorption capacity and regeneration of some novel polymer adsorbents

In recent years, research has been undertaken on new PAs that may be used for the adsorption of VOSCs (Jung et al., 2017; Noshadi et al., 2016; Jafari et al., 2015; Mito-oka et al., 2013). So far, the highest D4 adsorption capacity has been obtained by Jafari et al. (2015) on a newly synthesized polymer by the copolymerization of DVB with 1-vinylimidazole (Table 10.12). The new adsorbent, in addition to its large specific surface area, is unparalleled among known adsorbents for the large pore volume (Table 10.11). Moreover, the developed series of adsorbents (PDVB-VI-x) proved to be easily regenerable at relatively low temperatures (100°C), without losing more than 10% of the initial adsorption capacity after five cycles, regardless of gas humidity. This indicates the weak strength of D4 binding to the adsorbent surface and the lack of D4 polymerization, characteristic especially for AC. The ease of PA regeneration relative to other adsorbents has also been confirmed by Jung et al. (2017) and Noshadi et al. (2016). The former developed a novel polyacrylic acid-based polymer (RPA), showing that, >95% D5 desorption is possible already at 80°C. In turn, Noshadi et al. confirmed that, after 10 cycles, the adsorption capacity of D4 polymer P(DVB-ACAM) is still very high (~1000 g/kg).

Table 10.12

Kind of tested medium

Kind of AP

Kind of adsorbate

Adsorption capacity, g/kg


RS1

D4

300

PDVB-VI-0.12

2370a, 2360b

PDVB

1951a, 1940b

PDVB-VI-0.24

1586a, 1582b

PDVB-VI-0.50

1384a, 1381b

P(DVB-ACAM)

2220

D5


Noshadi et al. (2016)

500

VMS, volatile methylsiloxane; PAs, polymer adsorbents

aRH=0%.

bRH=50%.

The main advantages of the above-analyzed PAs in the area of VOSC removal from biogas lies in their very high adsorption capacity, ease of regeneration, and hydrophobicity. A significant drawback of PAs in relation to mineral and carbon adsorbents is, however, their high production costs.

10.2.4 Absorption methods

Gas–liquid absorption, that is, a diffusive process of physical and/or chemical mass exchange between the gas and liquid phase, is usually carried out in spray scrubbers or packed columns, ensuring an adequate mass exchange surface and phase time. Since most VOSCs poorly dissolve in water, nonvolatile, high boiling organic solvents are most often used for their physical absorption. Chemical absorption tests have also been undertaken by using concentrated mineral acids (including HNO3 and H2SO4) at elevated temperatures—up to 60°C (Schweigkofler and Niessner, 2001). Their task was to cleave Si-O bonds, leading to the formation of nonvolatile PDMS polymers. However, despite the high effectiveness of these methods, they have not been implemented—due to corrosion and environmental problems. Further research on the use of bases as VMS absorbents, due to the blocking installation with carbonates precipitated by their reaction with CO2, as well as tests of organic absorbents, such as n-

tetradecane, n-dodecane, n-hexadecane (Huppmann et al., 1996; Schweigkofler and Niessner, 2001), which are toxic, flammable, expensive, and require cooling to prevent their evaporation, were also abandoned. Only better water-soluble VOSCs, that is, TMS-OH and TMS (and partially L2 and D3), can be effectively removed from biogas by aqueous washing. This is a frequently practiced method of biogas pretreatment, mainly aimed at removing foam, solid particles, and parts of H2S, NH3, and CO2, as well as biogas cooling. The latter is desirable if the next stage of biogas purification is physical adsorption using a hydrophilic adsorbent or physical absorption in an organic solvent (as is known, the efficiency of both increases with a decrease in temperature). In the second case, an additional effect is to reduce absorbent losses through evaporation. Another way to improve VMS absorption can be by adding organic liquid to water-based absorbent (Popat and Deshusses, 2008). Only pressurized water scrubbing [e.g., at 2–2.5 MPa, 10°C–25°C, and pH in the range of 4.4–4.9—according to Läntelä et al. (2012)] can reduce VMSs by more than 99%. The absorption of VMSs in hydrocarbon oils was tested in the United Kingdom and (along with halides removal) but without promising results [according to EA (2010), ~60% removal was achieved]. This method does not seem promising, either due to fire and/or toxicological hazards. Practically, only Selexol, produced on the basis of polyethylene glycol dimethyl ether, found commercial use as an absorbent of VOSCs (as well as H2S, CO2, and water vapor) (Wheless and Pierce, 2004; EA, 2010). Its advantages include low vapor pressure, weak binding with absorbed gases, low affinity for methane, low viscosity, chemical stability, low freezing point, noncorrosivity, and nontoxicity. In addition, it can be easily regenerated using air stripping columns. Selexol is used in many comprehensive landfill gas purification plants in the United States, before introducing biogas into the natural gas grid (Arnold, 2009; EA, 2010). However, it is relatively expensive [~6.5 €/kg—according to Cormos et al. (2018)]. The Kryosol process is also described (EA, 2010), and uses chilled (~−70°C) methanol at a pressure of 2.8 MPa. Its main task is to absorb CO2 with other acid gases and water vapor, but VMSs also dissolve very well in it. Due to the energy consumption, as well as flammability and toxicity, use of the Kryosol process only for the removal of VOSCs does not seem justified.

Silicone oils are potentially good VOSC absorbents. These are liquids with low volatility and high thermal and chemical stability, that is, suitable for repeated regeneration. An example would be 47V20 oil used in laboratory studies for the absorption of L2, L3, and D4 by Ghorbel et al. (2014). The authors showed that the cyclic VMSs present more affinity with the oils than linear ones. The removal efficiency of D4 reached almost 100%, and for L2 and L3 it was ~61% and ~82%, respectively. The cost of oil, however, is quite significant (14 €/kg). Physical absorption, unlike chemical absorption, however, has the disadvantage, especially with more volatile VOSCs, that there is a risk of desorption of absorbed pollutants due to an increase in biogas temperature or flow rate— which is often the case with landfill gas. Further research is needed toward competitive, environmentally safe, and more effective VOSC absorbents and process optimization.

10.2.5 Membrane techniques

Membrane separation consists of the selective, diffusive permeation of a component or components of a mixture-feed, through a thin layer of a porous barrier [usually in the form of interconnected and hollow plastic fibers, for example, made of silicone polymers, cellulose acetate, polycarbonates, polyamides, polyimides, etc. (Basu et al., 2010; Chen et al., 2015; Scholz et al., 2013)], and/or through a liquid film, due to the difference in concentration and pressure on both sides of the membrane. This is a relatively new method in gas purification, but is rapidly gaining importance due to the increasing availability of selective and cheap polymeric materials. As a part of biogas purification technology, semipermeable membranes have been commercially used so far only for CO2/CH4 separation. Their use to remove VOSCs is still in research. For example, Ajhar et al. (2006, 2012) showed that the PDMS membrane has good H2O and VMS permeability. It could therefore be used for simultaneous drying of biogas. Simultaneously, some chlorinated hydrocarbons, often found in landfill gas, can be removed in this way.

The main disadvantage of the membrane technique may be the high energy demand to power the compressors or vacuum pumps necessary to produce required pressure difference [up to 0.8 MPa, according to (Petersson and Wellinger, 2009)]. In addition, it is necessary to preremove substances that may block the membrane (e.g., aerosols from the digester and oil vapors from the compressor) or damage its structure (e.g., acidic compounds and impurities that can polymerize on the membrane surface or cause precipitation). The high potential for VMS adsorption on various materials, as well as the risk of losing some of the CH4 (~7% according to Ajhar et al., 2012), may also be an obstacle to VOSC removal in this way. The important advantages of this method include relatively low investment costs (GTI, 2014), which are associated with low space requirements, high selectivity, possibility of continuous operation, as well as modularity and compactness of the plant. It would be particularly interesting and worthy of further research to combine VOSCs and H2S removal with this technique before introducing biogas into the natural gas grid—because such biogas requires compression anyway. The method produces neither sewage nor solid waste (apart from the possible periodic replacement of the membrane itself), so it can be environmentally competitive to the sorption techniques. However, as in the case of sorbent regeneration, there are no standardized and environmentally safe ways of neutralizing permeate with separated VOSCs and other impurities. This requires further research. A promising solution, increasing process selectivity and reducing the content of VOSCs in biogas, would be to combine membrane separation with absorption in one device, using, for example, the PDMS membrane and the solvent in the form of silicone oil on the permeate side.

10.2.6 Biological methods

Biological gas cleaning consists of two parallel or separable processes: sorption in water or into moist active surfaces of mineral or organic solids, inhabited by microorganisms, and biodegradation, that is, biochemical transformation of an absorbed pollutant, through intermediates, into CO2 and H2O, with the

participation of biocatalysts. The process can be realized using:

1. Bioscrubbers fed with an aqueous suspension of activated sludge (which then undergoes biological regeneration in a separate reactor) or with water (then directed to the biological stage)—this method is only suitable for water-soluble impurities, and therefore it is not proper for most VOSCs; 2. Biofilters filled with a porous and moist bed of crushed organic materials in the form of bark, peat, compost, straw, fertile soil, coconut fibers, etc., which are mixed with leavening agents (e.g., sawdust, wood chips, polystyrene shapes, nut shells, pomace), additives improving the structure of the bed (e.g., expanded clay, perlite) and, if necessary, additives correcting its pH (e.g. CaO); 3. Biotrickling filters (BTFs), that is, countercurrent drip columns, filled with ceramic or plastic packing, covered with the biofilm of a suspension of microorganisms dispersed in the upper part of the apparatus.

A precondition for the process is at least partial water solubility and biodegradability of impurities. In the case of VMSs, the solubility condition is poorly met (Table 10.1). However, this does not exclude the use of the aforementioned techniques (2) and (3). The bioreactor must, however, ensure a sufficient time and large sorption surface, which unfortunately significant affects investment costs. The same applies to the second necessary condition—biodegradability. According to EChA European Chemical Agency (2018), VMSs are not readily biodegradable in water at aerobic conditions. Observations carried out for 28 days showed that the degree of biodegradation ranged from 0% for L3, L4, and L5 to 4.5% for D6, at 20°C. Similarly, Accettola and Haberbauer (2005) showed that the degradation of VMSs under aerobic conditions (via Pseudomonas bacteria) is very slow. The authors proposed a flow diagram of a feasible D4 biodegradation process. The degradation of cyclic VMSs is described as a multistage hydrolysis process, beginning with the ring-opening hydrolysis (involves the hydroxylation of a methyl groups), proceeding through the formation of linear oligomeric siloxane diols, and ending with nonvolatile dimethylsilanediol (DMSD), silanetriols, and formaldehyde—as intermediate

biodegradation products. These, in turn, can be broken down—by aerobic bacteria of the Arthrobacter genus or by fungi Fusarium oxysporum—to silicic acids, methanol, and finally to CO2 and H2O (Soreanu et al., 2011). According to Sabourin et al. (1996), the degradation rate does not exceed 4%/month. This was confirmed by the report of Wang et al. (2013) who found the half-life of D5 in aquatic sediments under biotic aerobic and anaerobic conditions to be above 1000 days, at 24°C. In turn, the anaerobic decomposition of the cyclic VMSs in activated sludge was investigated by Xu et al. (2013). After 60 h of the process, they found that the degree of degradation was in the range of 44.4%–62.8% for D4 and D5 and in the range of 3.0%–18.1% for D3 and D6. Anaerobic respiration in this case involves the use of another electron acceptor (e.g., nitrates) instead of oxygen. Thus the poor biodegradability and high hydrophobicity of the main VMSs are the major limitations in the use of biological methods for their removal from biogas. The consequence is high mass transfer resistance in the liquid phase. However, absorption efficiency can be improved by using an additional, nonmiscible, and hardly biodegradable organic absorbent, for example, silicone oil or oleyl alcohol (Popat and Deshusses, 2008), improving the solubility of hydrophobic compounds. Meanwhile, the effectiveness of biodegradation, by adding specific enzymes (Brandstadt, 2005) or biosurfactants (Accettola et al., 2008), increased the availability of D4 to bacterial cells. So far the biodegradation of VMSs by laboratory BTFs has not led to satisfactory results. For example, the D4 removal efficiency of only 43% at an empty bed residence time of 19.5 min was achieved at aerobic conditions by Popat and Deshusses (2008). Such a long time is not feasible in practical applications due to the very large reactor size required. The authors noted similarly poor D4 removal under anaerobic conditions. Slightly more effective biological removal of D4 has been obtained by Li et al. (2014)—74% at a retention time of 13.2 min, using the strain of Pseudomonas aeruginosa. The authors stated that rhamnolipids biosurfactants produced by the aforementioned strain contributed to the improvement of D4 degradation efficiency. They identified DMSD, methanol, silicic acid, and CO2 as the end products of degradation and proposed a possible metabolic pathway for D4 degradation by P. aeruginosa. Similar efficiency of D4 degradation (60.2%/24 min), using BTF, but also other bacterial strains—Phyllobacterium myrsinacearum—was obtained by Wang et al. (2014). It is worth emphasizing that these strains originated from sewage sludge from a silicone-producing factory (and thus were initially adapted to organic silicon compounds), in contrast to earlier studies by Popat and

Deshusses (2008), who used activated sludge from municipal STP. However, despite the relatively low efficiency of biological degradation processes, which does not yet allow for their practical implementation, studies have shown that for microorganisms such as Pseudomonas (P. aeruginosa, P. fluorescens, P. putida), Agrobacterium (A. radiobacter), Arthrobacter, and Fusarium oxysporum, some of the organic silicon compounds can be the only carbon source. Pseudomonas bacteria, which are widespread in nature and can be easily inoculated from the STP-activated sludge, can be particularly useful. Some of them function well under anaerobic conditions, which can be a particularly useful feature in the case of biogas treatment. In addition, they can produce the aforementioned biosurfactants. However, these microorganisms need a relatively long time to adapt, that is, to adjust their metabolic pathways to the breakdown of these unnatural compounds, such as VOSCs. In summary, despite research on the biological degradation of silicone compounds that has been ongoing since the 1990s, technologies for the biological removal of VOSCs from biogas are still in the experimental phase, and the results of previous tests using BTF are generally unsuccessful. The obtained mass transfer coefficients are many times lower than in the case of biological removal of H2S. First, the problem of too-long biodegradation time of these compounds—for example, by addition of an appropriate enzymes—has not been solved so far. Improving gas–liquid mass transfer by ixtures of organic liquids and biosurfactants, as well as selecting more efficient bacterial strains and optimal process conditions, remain a challenge. There is also no research on additional nutrients, that is, sources of nitrogen, phosphorus, and other biogenic elements, as well as microelements that may be missing in biogas. And finally, there is no research testing this technology on real biogas. Many researchers rightly point out that simultaneous biological removal of VOSCs—at their low concentrations—and H2S under anaerobic conditions or with a small addition of oxygen could be a promising solution. An example of such a method may be the use of some ZEs for the simultaneous adsorption of these compounds and as a biofilter filling (Gaj, 2017). Thus due to the potential competitiveness of this technology (in of operational costs and environmental effects) to known physical and chemical methods, and the empirically confirmed possibility of VOSCs decomposition by selected microorganisms, research should be continued toward process intensification. There are hopes for the use of Methylibium sp. bacteria in the anaerobic process of simultaneous removal of VOSCs and other VOCs such as hexane and toluene

(Boada et al., 2020). It seems however, that this process can only proven and more effective adsorption methods.

10.3 Combined methods for volatile organic silicon compounds removal from biogas

In practice, various technical solutions for biogas processing are connected in series, thanks to which both the entire system and its individual stages are more efficient and less costly to operate. The most commonly used hybrid system is the combination of condensation drying—which also ensures partial removal of solid particles and gases soluble in condensate—with adsorption into AC or SG. Among emerging VOSC removal technologies, membrane techniques may be preferably combined with absorption or adsorption methods, and biological methods—with adsorption, for example, on ZEs, along with removal of H2S. In the case of biogas upgrading to biomethane, the above system can be supplemented with cryogenic separation (Chen et al., 2015). In addition, the membrane technology requires the initial removal of contaminants such as dust, H2O, H2S, NH3, and VOCs, for example, by condensation and AC adsorption, to prevent blocking or damage to the membrane. A promising method of biogas purification can be VOSCs adsorption using halloysite combined with the biological removal of H2S in a single device (Gaj, 2017). An interesting solution may be also a combination of VOSCs adsorption into AC, with their conversion into more water-soluble silanediols, which can then be biologically degraded (de Arespacochaga et al., 2019). Such solutions, in addition to providing more comprehensive purification, extend the lifetime of subsequent stages.

10.4 Comparison of the methods for reducing the content of volatile organic silicon compounds in biogas

Comparative analysis of the methods, including their performance parameters, favorable and unfavorable features, and cost comments is presented in Table 10.13.

Table 10.13

Technology

Removal efficiency and performance

1

2 Pretreatment methods

<90%

Condensation methods

<50%

Freezing methods—up to −30°C

<90%

Adsorption into activated carbon (AC)

<99% Real biogas: 5–63 g/kg. Synthetic biogas (N2)

Adsorption into conventional silica gel (SG)

<99% Synthetic biogas (N2): 17–259 g/kg

Adsorption into zeolites (ZEs)

<99% Synthetic biogas (N2): 11–276 g/kg

Adsorption into commercial alumina (AA)

<99% Synthetic biogas (N2): 127 g D4/kg

Adsorption into polymer adsorbents (PA)

<99% Synthetic biogas (N2): 500 g D5/kg, 300–2370

Absorption in water

<50%

Absorption in hydrocarbons, alcohols, and hydrocarbon oils

<95%

Absorption in Selexol

<95%

Absorption in silicone oil

<99% for D4, ~ 60% for L2, ~ 80% for L3

Membrane techniques

<80%<99% (when simultaneous absorption is used)

Biological methods using biotrickling filter (BTF)

In relation to D4: 43% at a residence time of 19.5 mi

10.5 Conclusions and future perspective

Due to the relatively low costs, availability, ease of use, and possibility for simultaneous removal of the other undesirable biogas components, that is, hydrogen sulfide and halides, adsorption using AC is still the most widely used technique for removing VOSCs from biogas. Its use seems to be particularly justified in the case of low concentrations of VOCSs (preferably below 1 mg/m³), for example, as the final stage of a comprehensive biogas purification. However, this technique has many disadvantages, the largest of which is the practical inability to regenerate the spent AC—due to the tendency for VMS polymerization on its surface. This necessitates frequent exchange of the bed and its utilization as a nuisance waste. Other drawbacks are the nonselective operation, the need for biogas pretreatment in of humidity, temperature, and possible removal of competing contaminants, and the risk of mutual displacement of VMSs, which may result in rapid breakthrough by lighter and more volatile ones. It is known that the adsorption activity increases with increasing specific surface area and the pore volume of the AC, and decreases with increasing gas humidity and temperature. However, the impact of the AC properties (porosity, pore size, grain size and shape, type of chemical used for activation or impregnation) and the impact of biogas composition on the adsorption efficiency of individual VOSCs and their possible interactions on the adsorbent surface, including polymerization and superseding, are still poorly understood. There are still no effective, cheap, and safe methods of spent AC regeneration and standards for its management as a waste. AC compared to SG and ZE has a larger or comparable specific surface area and pore volume, however it is less selective and more difficult for thermal regeneration. On the other hand, due to the wide range of pore sizes, it is the most universal adsorbent. In turn, the disadvantage of AC in biogas applications, compared to SG and ZE, is its greater CH4 adsorption capacity. SG has greater or comparable to AC adsorption capacity of VMSs, but biogas requires deep drying (RH<10%). Regeneration of SG is also problematic. SG is not suitable as an H2S adsorbent. In turn, ZEs seem to be better adsorbents for simultaneous removal of H2S and VOSCs (due to their alkalinity, hydrophobicity, and mechanical and thermal resistance), but also in the case that potential VMS polymerization is found.

Alumina and polymer resin adsorbents are much easier to regenerate. The latter additionally show a much larger total pore volume, adsorption capacity, and selectivity compared to the above. They usually have hydrophobic properties, which are beneficial for biogas applications. Although they are currently more expensive, they seem to be the most promising adsorbents. Physical absorption in organic solvents, like physical adsorption, is a nondestructive and regenerative method. It can ensure simultaneous drying of biogas and removal of sulfur and chlorine compounds, without significantly reducing the concentration of CH4. Advantageously, this process can be preceded by absorption in water, which, in addition to eliminating some of the impurities, allows cooling of the biogas, which increases the efficiency of the removal of VOSCs in the second stage. Physical absorption has the advantage over physical adsorption that it allows easier and more effective regeneration of the sorbent. Recently, more often, especially for biogas upgrading, the aforementioned processes are intensified using increased pressure. Apart from the well-tested Selexol, silicone oils seem to be the most promising absorbents. However, further research is needed on optimal absorption and regeneration conditions, as well as new apparatus designs for these processes. The use of membrane techniques to remove VOSCs from biogas is still in the research phase. However, with the increasing availability of new and cheap polymer materials, the chances of their full industrial implementation are increasing. Their use for biogas upgrading seems particularly justified due to the need for biogas drying (e.g., using PDMS membranes) and compression. Earlier, however, the problem of permeate utilization or neutralization should be solved. It is also worth undertaking research on the intensification of the process by combining it with absorption in the same apparatus. Biological neutralization of VOSCs is another technology with great development potential. Unfortunately, the hydrophobicity and poor biodegradability of most VMSs do not favor their intensive removal by this method. However, there are some ways to improve gas–liquid mass exchange by adding nonbiodegradable and immiscible organic solvent to water or activated sludge, and increasing the rate of VOSC biodegradation by adding specific biocatalysts and biosurfactants. The latter may be produced by some bacteria, so choosing the right strain seems to be crucial. However, previous attempts using BTF, due to the very long time required, did not give successful results. In addition to the development of the above aspects, further research should

focus on combining the biological removal of H2S (which has already found many full industrial applications), with the simultaneous removal of VOSCs, for example, using ZEs as the adsorbent and the habitat of microorganisms. Most methods of reducing the VOSC content in biogas, in particular physical sorption methods, but also membrane techniques, do not neutralize these compounds. They are only transferred to another environment, ending up in the air (as a result of physical sorbent regeneration or the membrane process) or in STPs and landfills, from where they are also released into the atmosphere. In the light of recent reports on the toxicity and bioaccumulation potential of some VMSs, this is problematic. In this respect, destructive methods that transform VOSCs into SiO2, CO2, and H2O, that is, biological or chemical oxidation methods, are future-proof. The latter can also lead, for example, through the action of OH radicals, to the transformation of VOSCs into substances more soluble in water and more easily biodegradable in the environment, that is, silanes and silanols. However, even using biological methods, the release of toxic intermediate decomposition products such as aldehydes or methanol cannot be excluded. Due to the ongoing climate change crisis and the need for intensive development of renewable energy sources, there is an obvious need for further research aimed at developing biogas technologies, including methods for removing the most troublesome biogas pollutants, such as VOSCs.

References

Accettola and Haberbauer, 2005 Accettola F, Haberbauer M. Control of siloxanes. In: Lens P, Westermann P, Habebauer M, Moreno A, eds. Biofuels for Fuel Cells, Renewable Energy from Biomass Fermentation. London: IWA Publishing; 2005;445–454. Accettola et al., 2008 Accettola F, Guebitz GM, Schoeftner R. Siloxane removal from biogas by biofiltration: biodegradation studies. Clean Techn Environ Policy. 2008;10:211–218. Ajhar and Melin, 2006 Ajhar M, Melin T. Siloxane removal with gas permeation membranes. Desalination. 2006;200:234–235. Ajhar et al., 2010 Ajhar M, Travesset M, Yuce S, Melin T. Siloxane removal from landfill and digester gas—a technology overview. Bioresour Technol. 2010;101:2913–2923. Ajhar et al., 2012 Ajhar M, Bannwarth S, Stollenwerk KH, et al. Siloxane removal using silicone-rubber membranes. Sep Purif Technol. 2012;89:234–244. Appels et al., 2008 Appels L, Baeyens J, Dewil R. Siloxane removal from biosolids by peroxidation. Energ Convers Manage. 2008;49(10):2859–2864. Arnold, 2009 Arnold, M., 2009. Reduction and monitoring of biogas trace compounds. VTT Tiedotteita—RN 2496, Finland. Arnold and Kajolinna, 2010 Arnold M, Kajolinna T. Development of on-line measurement techniques for siloxanes and other trace compounds in biogas. Waste Manage. 2010;30:1011–1017. Basu et al., 2010 Basu S, Khan AL, Cano-Odena A, Liu C, Vankelecom IFJ. Membrane-based technologies for biogas separations. Chem Soc Rev. 2010;39:750–768.

Boada et al., 2020 Boada E, Santos-Clotas E, Bertran S, et al. Potential use of Methylibium sp as a biodegradation tool in organosilicon and volatile compounds removal for biogas upgrading. Chemosphere. 2020;240:124908. Brandstadt, 2005 Brandstadt KF. Inspired by nature: an exploration of biocatalyzed siloxane bond formation and cleavage. Curr Opin Biotechnol. 2005;16(4):393–397. Cabrera-Codony et al., 2014 Cabrera-Codony A, Montes-Moran MA, SanchezPolo M, Martin MJ, Gonzalez-Olmos R. Biogas upgrading: optimal activated carbon properties for siloxane removal. Environ Sci Technol. 2014;48:7187– 7195. Cabrera-Codony et al., 2015 Cabrera-Codony A, Gonzalez-Olmos R, Martin MJ. Regeneration of siloxane-exhausted activated carbon by advanced oxidation processes. J Hazard Mater. 2015;285:501–508. Cabrera-Codony et al., 2017 Cabrera-Codony A, Georgi A, Gonzalez-Olmos R, Valdés H, Martín MJ. Zeolites as recyclable adsorbents/catalysts for biogas upgrading: removal of octamethylcyclotetrasiloxane. Chem Eng J. 2017;307:820–827. Cabrera-Codony et al., 2018 Cabrera-Codony A, Santos-Clotas E, Ania CO, Martín MJ. Competitive siloxane adsorption in multicomponent gas streams for biogas upgrading. Chem Eng J. 2018;344:565–573. Chen et al., 2015 Chen X, Vinh-Thang H, Ramirez AA, Rodrigue D, Kaliaguine S. Membrane gas separation technologies for biogas upgrading. RSC Adv. 2015;5:24399–24448. Cormos et al., 2018 Cormos A-M, Szima S, Fogarasi S, Cormos C-C. Economic assessments of hydrogen production processes based on natural gas reforming with carbon capture. Chem Eng Trans. 2018;70:1231–1236. de Arespacochaga et al., 2014 de Arespacochaga N, Valderrama C, Mesa C, Bouchy L, Cortina JL. Biogas deep clean-up based on adsorption technologies for solid oxide fuel cell applications. Chem Eng J. 2014;255:593–603. de Arespacochaga et al., 2015 de Arespacochaga N, Valderrama C, RaichMontiu J, Crest M, Mehta S, Cortina JL. Understanding the effects of the origin,

occurrence, monitoring, control, fate and removal of siloxanes on the energetic valorization of sewage biogas—a review. Renew Sust Energ Rev. 2015;52:366– 381. de Arespacochaga et al., 2019 de Arespacochaga N, Raich-Montiu J, Crest M, Cortina JL. Presence of siloxanes in sewage biogas and their impact on its energetic valorization. In: Homem V, Ratola N, eds. Volatile Methylsiloxanes in the Environment Book Series: The Handbook of Environmental Chemistry. Springer Nature Switzerland AG 2019. EA, 2010 EA. Guidance on Gas Treatment Technologies for Landfill Gas Engines UK: Environment Agency; 2010. EChA (European Chemical Agency), 2018 EChA (European Chemical Agency), 2018. Information on chemicals. https://echa.europa.eu/information-onchemicals/ed-substances (accessed 10.03.2018). EPRI, 2006 EPRI. Assessment of Fuel Gas Cleanup Systems for Waste Gas Fueled Power Generation California: Electric Power Research Institute; 2006; 1012763, Technical Update. Finocchio et al., 2009 Finocchio E, Montanari T, Garuti G, et al. Purification of biogases from siloxanes by adsorption: on the regenerability of activated carbon sorbents. Energ Fuel. 2009;23(8):4156–4159. Gaj, 2017 Gaj K. Applicability of selected methods and sorbents to simultaneous removal of siloxanes and other impurities from biogas. Clean Techn Environ Policy. 2017;19(9):2181–2189. Gaj, 2018 Gaj K. Properties, toxicity and transformations of VMSs in the environment. In: Homem V, Ratola N, eds. Volatile Methylsiloxanes in the Environment Book Series: The Handbook of Environmental Chemistry. Springer Nature Switzerland AG 2018. Gaj and Pakuluk, 2015 Gaj K, Pakuluk A. Volatile methyl siloxanes as a potential hazardous air pollutants. Pol J Environ Stud. 2015;24(3):937–943. Ghorbel et al., 2014 Ghorbel L, Tatin R, Couvert A. Relevance of an organic solvent for absorption of siloxanes. Environ Technol. 2014;35(1–4):372–382.

Giraudet et al., 2014 Giraudet S, Boulinguiez B, Cloirec PL. Adsorption and electrothermal desorption of volatile organic compounds and siloxanes onto an activated carbon fiber cloth for biogas purification. Energy Fuels. 2014;28:3924– 3932. Gislon et al., 2013 Gislon P, Galli S, Monteleone G. Siloxanes removal from biogas by high surface area adsorbents. Waste Manage. 2013;33:2687–2693. Gong et al., 2015 Gong H, Chen Z, Fan Y, Zhang M, Wu W, Wang W. Surface modification of activated carbon for siloxane adsorption. Renew Energy. 2015;83:144–150. GTI, 2014 GTI, 2014. Conduct a Nationwide Survey of Biogas Cleanup Technologies and Costs, Final Report, AQMD Contract #: 13432, Gas Technology Institute, Illinois. Hepburn et al., 2015 Hepburn CA, Martin BD, Simms N, McAdam EJ. Characterization of full-scale carbon ors for siloxane removal from biogas using online Fourier transform infrared spectroscopy. Environ Technol. 2015;36(2):178–187. Huppmann et al., 1996 Huppmann R, Lohoff HW, Schröder HF. Cyclic siloxanes in the biological waste water treatment process—determination, quantification and possibilities of elimination, Fresenius. J Anal Chem. 1996;354:66–71. Jafari et al., 2015 Jafari T, Noshadi I, Khakpash N, Suib SL. Superhydrophobic and stable mesoporous polymeric adsorbent for siloxane removal: D4 superadsorbent. J Mater Chem A. 2015;3:5023–5030. Jiang et al., 2016 Jiang T, Zhong W, Jafari T, et al. Siloxane D4 adsorption by mesoporous aluminosilicates. Chem Eng J. 2016;289:356–364. Jung et al., 2017 Jung H, Lee D–Y, Jurng J. Low-temperature regeneration of novel polymeric adsorbent on decamethylcyclopentasiloxane (D5) removal for cost-effective purification of biogases from siloxane. Renew Energy. 2017;111:718–723. Kajolinna et al., 2015 Kajolinna T, Aakko-Saksa P, Roine J, Kåll L. Efficiency testing of three biogas siloxane removal systems in the presence of D5, D6, limonene and toluene. Fuel Process Technol. 2015;139:242–247.

Kazuyuki et al., 2007 Kazuyuki O, Masaki T, Tadao M, Hiroshi K, Nobuo T, Akira K. Behavior of siloxanes in a municipal sewage-treatment plant. J Jpn Sew Work Assoc. 2007;44(531):125–138. Läntelä et al., 2012 Läntelä J, Rasi S, Lehfinen J, Rintala J. Landfill gas upgrading with pilot scale water scrubber: performance assessment with absorption water recycling. Appl Energy. 2012;92:307–314. Lee et al., 2001 Lee S-H, Cho W, Song T-Y, et al. Removal process for octamethylcyclotetrasiloxane from biogas in sewage treatment plant. J Ind Eng Chem. 2001;7(5):276–280. Li et al., 2014 Li Y, Zhang W, Xu J. Siloxanes removal from biogas by a labscale biotrickling filter inoculated with Pseudomonas aeruginosa S240. J Hazard Mater. 2014;275:175–184. Liu et al., 2019 Liu Y-H, Meng Z-Y, Wang J-Y, Dong Y-F, Ma Z-C. Removal of siloxanes from biogas using acetylated silica gel as adsorbent. Pet Sci. 2019;16:920–928. Matsui and Imamura, 2010 Matsui T, Imamura S. Removal of siloxane from digestion gas of sewage sludge. Bioresour Technol. 2010;101(1):S29–S32. Miguel et al., 2002 Miguel GS, Lambert SD, Graham JD. Thermal regeneration of granular activated carbons using inert atmospheric conditions. Environ Technol. 2002;23(12):1337–1346. Mito-oka et al., 2013 Mito-oka Y, Horike S, Nishitani Y, et al. Siloxane D4 capture by hydrophobic microporous materials. J Mater Chem A. 2013;1:7885– 7888. Montanari et al., 2010 Montanari T, Finocchio E, Bozzano I, et al. Purification of landfill biogases from siloxanes by adsorption: a study of silica and 13X zeolite adsorbents on hexamethylcyclotrisiloxane separation. Chem Eng J. 2010;165:859–863. Nam et al., 2013 Nam S, Namkoong W, Kang J, Park J, Lee N. Adsorption characteristics of siloxanes in landfill gas by the adsorption equilibrium test. Waste Manage. 2013;33(10):2091–2098.

Noshadi et al., 2016 Noshadi I, Kanjilal B, Jafari T, et al. Hydrophobic mesoporous adsorbent based on cyclic amine–divinylbenzene copolymer for highly efficient siloxane removal. RSC Adv. 2016;6:77310–77320. Ortega and Subrenat, 2009 Ortega DR, Subrenat A. Siloxane treatment by adsorption into porous materials. Environ Technol. 2009;30(10):1073–1083. Oshita et al., 2010 Oshita K, Ishihara Y, Takaoka M, et al. Behaviour and adsorptive removal of siloxanes in sewage sludge biogas. Water Sci Technol. 2010;61(8):2003–2012. Oshita et al., 2014 Oshita K, Omori K, Takaoka M, Mizuno T. Removal of siloxanes in sewage sludge by thermal treatment with gas stripping. Energ Conv Manage. 2014;81:290–297. Oshita et al., 2015 Oshita K, Fujime M, Takaoka M, Fujimori T, Appels L, Dewil R. Siloxane removal and sludge disintegration using thermo-alkaline treatments with air stripping prior to anaerobic sludge digestion. Energ Conv Manage. 2015;96:384–391. Petersson and Wellinger, 2009 Petersson A, Wellinger A. Biogas upgrading technologies – developments and innovations Task 37 - Energy from biogas and landfill gas IEA Bioenergy 2009. Popat and Deshusses, 2008 Popat SC, Deshusses MA. Biological removal of siloxanes from landfill and digester gases: opportunities and challenges. Environ Sci Technol. 2008;42(22):8510–8515. Prabucki et al., 2001 Prabucki, M.-J., Doczyck, W., Asmus, D., 2001. Removal of organic silicon compounds from landfill and sewer gas. In: Proceedings Sardinia 2001. VIII Int. Waste Management and Landfill Symp. Vol. II, 631– 639. Repelewicz and Choma, 2006 Repelewicz M, Choma J. Fizykochemiczne właściwości niemodyfikowanych i chemicznie modyfikowanych węgli aktywnych na przykładzie węgla WG-12. In: Dębowski Z, ed. Węgiel aktywny w ochronie środowiska i przemyśle. Politechnika Częstochowska 2006;169–180. Ryckebosch et al., 2011 Ryckebosch E, Drouillon M, Vervaeren H. Techniques for transformation of biogas to biomethane. Biomass Bioenerg.

2011;35(5):1633–1645. Sabourin et al., 1996 Sabourin CL, Carpenter JC, Leib TK, Spivack JL. Biodegradation of dimethylsilanediol in soils. Appl Environ Microbiol. 1996;62(12):4352–4360. Scholz et al., 2013 Scholz M, Melin T, Wessling M. Transforming biogas into biomethane using membrane technology. Renew Sust Energ Rev. 2013;17:199– 212. Schweigkofler and Niessner, 2001 Schweigkofler M, Niessner R. Removal of siloxanes in biogases. J Hazard Mater. 2001;83(3):183–196. Sigot et al., 2014 Sigot L, Ducom G, Benadda B, Labouré C. Adsorption of octamethylcyclotetrasiloxane on silica gel for biogas purification. Fuel. 2014;135:205–209. Sigot et al., 2015 Sigot L, Ducom G, Germain P. Adsorption of octamethylcyclotetrasiloxane (D4) on silica gel (SG): retention mechanism. Micropor Mesopor Mat. 2015;213:118–124. Sigot et al., 2016 Sigot L, Ducom G, Benadda B, Labouré C. Comparison of adsorbents for H2S and D4 removal for biogas conversion in a solid oxide fuel cel. Environ Technol. 2016;37(1):86–95. Soreanu et al., 2011 Soreanu G, Beland M, Falletta P, et al. Approaches concerning siloxane removal from biogas—a review. Can Biosyst Eng. 2011;53:8.1–8.18. Urban et al., 2009 Urban W, Lohmann H, Gomez JIS. Catalytically upgraded landfill gas as a cost-effective alternative for fuel cells. J Power Sources. 2009;193(1):359–366. Vagenknechtová et al., 2017 Vagenknechtová A, Ciahotný K, Vrbová V. Siloxanes removal from biogas using activated carbon. Acta Polytech. 2017;57(2):131–138. Wang et al., 2013 Wang D-G, Norwood W, Alaee M, Byer JD, Brimble S. Review of recent advances in research on the toxicity, detection, occurrence and fate of cyclic volatile methyl siloxanes in the environment. Chemosphere.

2013;93(5):711–725. Wang et al., 2014 Wang J, Zhang W, Xu J, Li Y, Xu X. Octamethylcyclotetrasiloxane removal using an isolated bacterial strain in the biotrickling filter. Biochem Eng J. 2014;91:46–52. Wheless and Gary, 2002 Wheless, E., Gary, D., 2002. Siloxanes in landfill and digester gas. Proceedings from the Solid Waste Association of North America 25th Annual Landfill Gas Symposium, 29–41. Wheless and Pierce, 2004 Wheless, E., Pierce, J., 2004. Siloxanes in landfill and digester gas update, SWANA 27th LFG Conference, SCS Energy, San Antonio, California. Xu et al., 2013 Xu L, Shi Y, Cai Y. Occurrence and fate of volatile siloxanes in a municipal wastewater treatment plant of Beijing, China. Water Res. 2013;47(2):715–724. Yang, 2003 Yang RT. Adsorbents: Fundamentals and Applications Hoboken, New Jersey: Wiley; 2003. Yu et al., 2013 Yu M, Gong H, Chen Z, Zhang M. Adsorption characteristics of activated carbon for siloxanes. J Environ Chem Eng. 2013;1:1182–1187. Zamorska-Wojdyła et al., 2012 Zamorska-Wojdyła D, Gaj K, Hołtra A, Sitarska M. Quality evaluation of biogas and selected methods of its analysis. Ecol Chem Eng S. 2012;19(1):77–87. Zhang et al., 2016 Zhang C, Jin M, Zhang S, Li Z, Li H. Removal of cyclic and linear siloxanes in effluents from a cosmetic wastewater treatment plant by electrochemical oxidation. Int J Electrochem Sci. 2016;11:6914–6921. Zhong et al., 2017 Zhong W, Jiang T, Jafari T, et al. Modified inverse micelle synthesis for mesoporous alumina with a high D4 siloxane adsorption capacity. Micropor Mesopor Mat. 2017;239:328–335.

Part III Biological upgrading systems

Outline

Chapter 11 Technologies for removal of hydrogen sulfide (H2S) from biogas

Chapter 12 Biological upgrading of biogas through CO2 conversion to CH4

Chapter 13 Bioelectrochemical systems for biogas upgrading and biomethane production

Chapter 14 Photosynthetic biogas upgrading: an attractive biological technology for biogas upgrading

Chapter 11

Technologies for removal of hydrogen sulfide (H2S) from biogas

Anish Ghimire¹, Raju Gyawali², Piet N.L. Lens³ and Sunil Prasad Lohani⁴, ¹1Department of Environmental Science and Engineering, Kathmandu University, Dhulikhel, Nepal, ²2Nepal Electricity Authority, Government of Nepal, Kathmandu, Nepal, ³3UNESCO—IHE Institute for Water Education, Delft, The Netherlands, ⁴4Department of Mechanical Engineering, Kathmandu University, Dhulikhel, Nepal

Abstract

H2S is produced in the anaerobic digestion process by the degradation of organic compounds (i.e., proteins) and reduction of the inorganic species (SO4²−) present in the feedstock. H2S is required to be removed from the biogas due to concerns about health, safety, and corrosion during transmission, storage, and use. Higher concentrations of H2S in biogas limit its applications in various technologies like turbines, internal combustion engines, fuel cells, and grid injection. This chapter reviews the range of different technological options, that is, physicochemical processes such as water scrubbing, membrane separation, adsorption onto activated charcoal, metal oxides, and absorption in organic solvents. Likewise, various in situ and ex situ biological biogas desulfurization methods such as the use of biological air filtration, direct and indirect microaeration techniques, and combined two-step chemical-biological processes are discussed. The prospects and limitations of the existing technologies are highlighted along with the outlooks for novel technologies such as the use of microalgae for biogas clean-up.

Keywords

Biogas upgrading; desulfurization; biogas purification; biomethane; biogas sweetening

Chapter outline

Outline


11.2 Technologies for removal of biogas contaminants 297

11.3 Physicochemical removal technologies 298

11.3.1 Absorption process 298 11.3.2 Adsorption process 300 11.3.3 Membrane separation 304

11.4 Ex situ removal using sulfur-oxidizing microorganisms 306

11.4.1 Biological air filtration 306

11.4.2 Microalgal removal of H2S 311

11.5 In situ H2S removal 312

11.5.1 In situ microaeration 312 11.5.2 Dosing iron salts/oxides into the digester 313

11.6 Combined chemical-biological processes 314

11.7 Comparison of H2S removal techniques 314

11.8 Conclusions 316

References 317

11.1 Introduction

Hydrogen sulfide is formed during the anaerobic digestion of organic waste and is present at a range of 100 to 10,000 ppm in biogas depending on the organic waste type (Zulkefli et al., 2016). The biogas composition and the concentration of impurities are affected by the type of anaerobic digestion system utilizing various feedstocks (Rasi et al., 2007). H2S is required to be removed from biogas due to concerns about health, safety, and corrosion. H2S present in biogas could lead to a deterioration of the equipment involved in production, conveyance systems, as well as inefficient operation or breakdown of the energy conversion systems (Zulkefli et al., 2016; Awe et al., 2017). The tolerance limits for application of biogas in various technologies are different (Rasi et al., 2007). The tolerance limits of H2S for fuels cells are stringent (<1 ppm) (Rasi et al., 2007). In addition, several European countries have standards for biogas quality as vehicle fuel and grid injection (Awe et al., 2017). The recommended levels of H2S in biogas for combustion are in the range 0.02%–0.05% wt./wt. (200 to 300 ppm) (Khoshnevisan et al., 2017). H2S is responsible for the deterioration of materials due to biogenic corrosion through the action of sulfur-oxidizing microorganisms depending on the environmental conditions (Gutiérrez-Padilla et al., 2010). In addition, H2S is oxidized to acidic sulfur dioxide (SO2) during combustion, which can be extremely corrosive to metal surfaces. The corrosion process speeds up at high temperatures, such as in the cylinder liner or piston ring of internal combustion engines. In an aqueous environment, metals and H2S interactions result in the formation of metal sulfide corrosion products that consequently lead to the potential for damage. The chemistry of sulfidic corrosion of iron is given by Eqs. 11.1–11.4 (Zheng et al., 2014; Salas et al., 2012). H2S dissolves in water to make an aqueous H2S solution, which is a mild acid that partly dissociates in two steps (Eq. 11.1). H2S attacks the iron, forming an unstable FeS layer. Under acidic conditions, the FeS is removed from the surface forming H2S again and enhancing corrosion.

(11.1)

(11.2)

(11.3)

In addition, H2S reduces the Fe³+ present in rust,

(11.4)

In addition to the corrosive nature of H2S, it causes poisoning of catalysts during steam reforming and must thus be removed from the gas stream. Considering the public health concerns of exposure to H2S, the National Institute for Occupational Safety and Health has a recommended exposure limit of 10 ppm (as a 10-minute ceiling) and Occupational Safety and Health Standards has a permissible exposure limit of 20 ppm (General Industrial Ceiling Limit) (OSHA, 2020). This chapter summarizes the state of the art of physical, chemical, and biological technologies for the removal of biogas contaminants with a focus on H2S removal. The prospects and limitations of the technologies are highlighted in this chapter. Novel technologies like the use of microalgae for the removal of

contaminants from biogas are also discussed.

11.2 Technologies for removal of biogas contaminants

Removal of H2S is usually carried out during the biogas cleaning or upgrading process, it may also be done directly in the digester through an oxidation process. The selection of the H2S removal process largely depends on various parameters, including (1) biogas and pollutant composition, (2) degree of pollutant elimination required, (3) required perm-selectivity (preferential permeation) of acid gas removal, (4) pressure, (5) temperature, (6) volumetric gas composition, (7) CO2/H2S ratio, and (8) desirability of sulfur recovery (Rezakazemi et al., 2017). Fig. 11.1 summarizes the available technologies for the ex situ (A) and in situ (B) removal of contaminants from biogas, especially H2S. In the ex situ processes, the biogas is upgraded in a separate unit process involving physical, chemical, or biological systems, while in situ biological H2S removal uses naturally occurring microorganisms in the anaerobic digesters. The physical and chemical processes are energy and chemical intensive. Therefore there is growing interest in the application of biotechnological processes for biogas upgrading as these are cheaper and environmentally friendly (Awe et al., 2017; Khoshnevisan et al., 2017; Dumont, 2015). However, the selection of optimal treatment methods depends on the initial biogas composition, potential application, treated gas specifications, and economic considerations. Several aspects of various H2S removal technologies are discussed herein.

Figure 11.1 (A) Ex situ and (B) in situ removal of contaminants (i.e., H2S) from biogas.

11.3 Physicochemical removal technologies

In this section, physical removal of H2S using water, organic solvents and metal oxides, pressure swing adsorption (PSA) and membrane separation systems are discussed as part of the physicochemical removal of H2S. The most common physical absorption technologies used for biogas cleaning and upgrading are water scrubbing, which had a 41% market share in 2012, followed by chemical scrubbing (22% market share) (Thrän et al., 2014). Due to its simplicity and effectiveness, physicochemical absorption is the most common technology for simultaneous H2S and CO2 removal (Thrän et al., 2014).

11.3.1 Absorption process

11.3.1.1 Water scrubbing

In this process, the simultaneous separation of CO2 and H2S from biogas occurs due to their higher solubility in water than CH4. In the normal operation of a water scrubber, the biogas is pressurized to about 6–10 bar and 40°C, and injected into the absorption column from the bottom side while a jet of water is sprayed from the top as shown in Fig. 11.2 (Bauer et al., 2013; Awe et al., 2017). The absorption column is mostly filled with packing material to increase the gas–liquid mass transfer and the desulfurized and upgraded biomethane is released from the upper side of the scrubber, while the liquid containing high concentrations of CO2 and H2S as well as a small amount of CH4 absorbed in water is expanded into a flash column with 2.5–3.5 bar pressure to minimize the methane loss (Fig. 11.2) (Bauer et al., 2013). The water can be reused depending on the scrubber type used. There are in general two types of commercially

available scrubbers in use: single- scrubbing and regenerative absorption.

Figure 11.2 Schematic of water scrubbing system for H2S removal.

In a single- scrubber, process water cannot be reused, and treated water is usually taken from a sewage treatment plant. In a regenerative absorption scrubber, the water can be regenerated in a desorption column while expanding at atmospheric pressure. The desorption process is carried out in a stripping tank with air blown (air sparging) when the H2S concentration is small. For biogas with high H2S concentrations, desorption is carried out by supplying steam or inert gas to avoid the formation of elemental sulfur by air stripping that may lead to operational problems (Ryckebosch et al., 2011; Angelidaki et al., 2018). The water regeneration is critical in the water scrubbing process because of the large volume of water required: about 200 m³/h water is needed to upgrade 1000 Nm³/g biogas (Bauer et al., 2013).

11.3.1.2 Physical absorption by using organic solvents

Physical absorption using organic solvents works under the same principle as water scrubbing. However, the absorbents for CO2 and H2S are different: organic solvent and water, respectively. In general, the organic solvents used for absorbing CO2 and H2S are mixtures of methanol and dimethyl ethers of polyethylene glycol. These are commercially available under the trade names, respectively, of Selexol and Genosorb. The affinity of H2S in these solvents is very high and their regeneration requires high temperatures to separate H2S from the solvents. In this process, the biogas is initially compressed to 7–8 bars and cooled to about 20°C before being supplied to the absorption column (Rezakazemi et al., 2017). The organic solvent is also cooled down before its addition, however, it is heated up to 80°C and expanded to about 1 bar in the desorption column to regenerate it (Bauer et al., 2013; Sun et al., 2015).

11.3.2 Adsorption process

11.3.2.1 Adsorption

Adsorption is the adhesion of atoms, ions, and molecules from a gaseous or liquid medium onto the surface of an adsorbent (Haycarb, 2019). Due to its simplicity and effectiveness, adsorption is considered as one of the most competitive technologies for desulfurization (>99%) (Zicari, 2003). Materials with a high surface area per unit weight are considered to have a good adsorption capacity. The pores having a diameter larger than the molecular diameter of the adsorbate are readily available for adsorption (Magomnang and Villanueva, 2014). There are two types of adsorption processes: physical, where the molecules are held in the pores not only by relatively weak van der Waals forces but also dipole–dipole forces, or dipole-induced forces and chemical, in which molecules are held by chemical bonding (Allegue and Hinge, 2014). The most commonly used adsorption methods for H2S removal from biogas are adsorption onto activated carbon and iron oxides (Allegue and Hinge, 2014). Adsorption is suitable for airstreams having low pollutant concentrations between 0.1–8 g/m³ and flow rates between 10–10,000 m³/h. Adsorbents can be regenerated by the use of steam or hot air. However, recovery of compounds is expensive, and the spent adsorbents are therefore often discarded to landfills or incinerated (Shareefdeen et al., 2005).

11.3.2.2 Adsorption onto activated carbon

Activated carbon is a commonly used carbonaceous, highly porous adsorptive medium (Haycarb, 2019). It has a complex structure, composed primarily of carbon atoms. The intrinsic pore network in the lattice structure of activated carbon allows the removal of impurities from gaseous and liquid media through

adsorption (Haycarb, 2019). Activated carbons are manufactured from carbonaceous materials such as coconut shell, peat, hard and soft wood, lignite coal, bituminous coal, or olive pits (Haycarb, 2019). The H2S removal ability of activated carbon depends on many factors such as the presence of pores of certain sizes, an affinity for water, and the chemical environment in the micropores (Adib et al., 1999). If a proper combination of surface chemistry (surface pH, acidic, and basic sites) and porosity is achieved, then activated carbons can act as excellent H2S adsorbent. An acidic environment boosts the formation of sulfur oxides because, in acidic environments, the concentration of hydrogen sulfide is low. These ions, when adsorbed, are susceptible to oxidation and converted to dispersed sulfur and then sulfur oxides (SO2 and SO3). In a basic environment (the concentration of HS− is higher) the sulfur atoms are brought together to be able to create polysulfides, which then polymerize to stable chains or form elemental sulfur (Bandosz, 2002). In of pH, its value should be higher than 5 to ensure the effectiveness of the H2S removal (the right concentration of HS− leads to oxidation of hydrogen sulfide to elemental sulfur) from the gas phase (Bandosz, 2002). A sufficient amount of water preadsorbed on the carbon surface facilitates the H2S dissociation process. Therefore the H2S oxidation rate increases significantly when applying an air stream with high relative humidity to the activated carbon (Kaliva and Smith, 1983). Another beneficial technique is the pretreatment of biogas using dry treated sludge as adsorbent before treating with activated carbon, resulting in a longer operational life of activated carbon and a more cost-effective purification process (Ortiz et al., 2014). This avoids the use of activated carbon for biogas with low H2S concentrations. The rate of H2S removal and the total load can be increased by impregnation or doping (preaddition of a reactive species) of the activated carbon with permanganate or potassium iodide (KI), potassium carbonate (K2CO3), or zinc oxide (ZnO) as catalyzers. H2S removal using ZnOimpregnated carbon is extremely efficient with final H2S concentrations <1 ppm (Petersson and Wellinger, 2009). The saturated activated carbons are usually replaced by fresh ones and they are either landfilled or incinerated for regeneration, which is an expensive process. The recovery of activated carbon depends on various parameters and their characteristics (San Miguel et al., 2001).

11.3.2.3 Adsorption on metal oxides

The use of metal oxides, for instance, iron oxide, as an adsorbent is also one of the common methods of H2S removal. It removes H2S from the gas phase by the formation of insoluble iron sulfides. H2S readily reacts with iron hydroxides or oxides to form iron sulfide. This reaction is slightly endothermic and is optimal at 25°C–50°C. The biogas to be purified must not be too dry, as moisture plays an important role during the reaction with iron oxide, but condensation should be prevented as the iron oxide material (e.g., pellets or grains) will stick together with water, which causes a reduction in the reactive surface (Wellinger and Lindberg, 2007). This method is also popular due to the advantage of regenerability of iron oxide through the elemental sulfur formation by the oxidation of iron sulfides with air. The regeneration process is highly exothermic, resulting in the possibility of self-ignition of the bed. However, regeneration is only possible for a limited number of times (Allegue and Hinge, 2014; Wellinger and Lindberg, 2007). Eqs. 11.5 and 11.6 show the adsorption reaction with metal oxide during purification and Eqs. 11.7 and 11.8 show the regeneration reactions of metal oxide. Purification:

(11.5)

(11.6)

Regeneration:

(11.7)

(11.8)

In iron sponge processes, an iron-impregnated adsorbent bed is formed by the deposition of hydrated iron oxides on the surface of wooden shavings. Oxides of other metals, like manganese, cobalt, nickel, lead, copper, and zinc also show similar behavior (Cebula, 2009). Iron oxides can be used mainly in two forms:

• Iron oxide pellets These are popularly used because the highest surface-to-volume ratios are achieved with pellets made of red mud, a waste product from aluminum production. However, the constraints of iron pellets are that their density is much higher than that of the wood chips and they are also quite expensive. At high hydrogen sulfide concentrations between 1000 ppm and 4000 ppm, 100 grams of pellets can bind 50 grams of sulfide (Wellinger and Lindberg, 2007; Krich et al., 2005). This removal process is popular in German and Swiss sewage treatment plants (Wellinger and Lindberg, 2007; Krich et al., 2005). • Iron oxide wood chips (iron sponge) These are wood chips covered with iron oxide. Being simple and requiring low capital cost, this technology has been widely adopted in the United States for the removal of H2S from biogas (Wellinger and Lindberg, 2007). The main advantage of these chips is the overall reduction in density due to the use of wood causing the surface-to-weight ratio to be very small. Approximately 20 grams of hydrogen sulfide can be adsorbed per 100 grams of iron oxide chips, which means it is also a cost-effective product (Wellinger and Lindberg, 2007).

However, care needs to be taken that the temperature does not rise too high while regenerating the iron wood chip filter (Wellinger and Lindberg, 2007).

11.3.2.4 Pressure swing adsorption system

PSA is a gas separation technology consisting of the selective adsorption of a gas onto adsorbent material, for instance a sieve, activated carbon, zeolite, or other materials with a high surface area, which can selectively adsorb and desorb the gas depending on the operating pressure (Augelletti et al., 2017). Fig. 11.3 shows a schematic of a PSA system. The main working principle of PSA technology is the affinity of pressurized gases to be attracted to the solid absorbent surfaces. In general, this process includes four steps, namely adsorption, blow-down, purge, and pressurization (Augelletti et al., 2017). When the compressed biogas of about 4–10 bars is supplied into the adsorption column, the adsorbent material selectively retains gases such as CO2, N2, O2, H2O, and H2S, while CH4 is able to through it (Bauer et al., 2013). The selectivity of a gas in the adsorbent highly depends on the equilibrium capacity (the relative equilibrium of each component of gases adsorbed in the given conditions) and/or the adsorption kinetics (differences in diffusion rates of different gases), thus making it a versatile technology for the separation of a gas from a gas mixture by adjusting the adsorbent material, its amount, residence time, and operating conditions (Medrano et al., 2019).

Figure 11.3 Schematic of a pressure swing adsorption system for H2S removal from biogas.

The adsorbent material can be regenerated by a desorption process, however, the adsorption of H2S is normally irreversible and thus it uses activated carbon dotted with potassium iodide (KI) for the H2S removal. This is referred to as impregnated activated carbon, which is a better adsorbent. The H2S is catalytically converted to elemental sulfur and water, which works similar to biological filters (i.e., in the presence of air) and the elemental sulfur is adsorbed on the pores of the activated carbon. The optimum pressure and temperature range for this process is 7–8 bar and 50°C–70°C, respectively, where the gas temperature is easy to achieve through the exothermic reaction which occurs during compression. The PSA system is designed as a regenerative system if the H2S concentration exceeds 3000 ppm in the gas phase (Petersson and Wellinger, 2009).

11.3.3 Membrane separation

11.3.3.1 Separation types

Membrane filtration is generally not used for selective removal of H2S from biogas, but becoming attractive to upgrade biogas to natural gas standards because of the reduced capital investment, ease of operation, low environmental impact, high reliability, low operation cost, ideal for remote locations, and adaptability (Zicari, 2003). Membrane separation is based on the selective permeability of the membranes, allowing the separation of the different biogas components (CH4, CO2, and H2S) (Awe et al., 2017). The separation process can be gas–gas (gas phase on both sides of the membranes) or gas–liquid (liquid absorbs the H2S and CO2 molecules diffusing through the membranes)

separation. In the gas–gas separation process, the membrane used is mainly polymeric, that is, cellulose acetate, polysulfone, poly(ether block amide) or PEBAX, and polyimide have been widely used for several commercial applications (Baker, 2012). In these polymeric membranes the diffusion coefficient and solubility of CO2, H2S, H2O, and O2 are higher than those of CH4. This results in a higher permeability of these gases, which through the membrane to the permeate side, while retaining CH4 at the inlet side. The difference in the gas/gas separation from the gas/liquid separation is that the liquid solution can be pure water, NaOH, and/or monoethanolamine solution, and the system is highly selective compared to solid membrane systems (Martin, 2008). In this separation process, the material used is a microporous hydrophobic membrane, which separates the gaseous stream flowing in one direction with the liquid stream flowing countercurrently that diffuses through the membrane (Martin, 2008). The amine solution can be regenerated by heating to release the absorbed gases that can be collected separately. Membrane processes work at high (>20–40 bar) or low (8–10 bar) pressure and the CH4-rich gas (92%–98%) remains to the side of the membrane with higher pressure, while all other gases (with 5%–10% CH4) permeate to the lower pressure side (Bauer et al., 2013). Depending on the required purity of methane, the membrane size can be larger or several membranes must be placed in series. Fig. 11.4 shows a schematic of the single-stage and two-stage membrane filtration processes. Losses of 5%–10% methane are possible as the CH4 may also through the membrane in order to achieve a higher purity of upgraded biogas (Martin, 2008; Kameyama et al., 1983). Therefore there is a need to balance the desire for high-purity methane and low methane loss. Recirculation of off-gases can help in reducing this loss. Single-stage systems give lower performance in of CH4 purity and recovery compared to multistaged systems (CH4 purity >99% and recovery >95%).

Figure 11.4 Schematic diagram of membrane filtration of biogas for the removal of H2S.

11.3.3.2 Membrane types

There are two types of membranes: solid membranes and immobilized liquid membranes. Solid membranes contain small pores and acts as a filter, whereas immobilized liquid membranes are pore-free, working on the principle of diffusion (Qi and Cussler, 1985). Thus, selectivity of membranes is controlled by solute size for solid membranes and by solubility for liquid membranes. An example of the transport of H2S in an immobilized liquid membrane consisting of aqueous carbonate solutions suggested by Matson et al. (1977) is as follows:

1. H2S dissolves in the liquid membrane (aqueous carbonate solutions) at the high-pressure side of the membrane; 2. H2S decomposes near the high-pressure side of the film by the reaction H2S → HS− + H+; 3. H+ produced in step (2) is consumed by CO3²− that is diffusing toward the high-pressure side by the reaction H+ + CO3²−→ HCO3−; 4. HS− and HCO3− diffuse from the high-pressure to the low-pressure side of the film; 5. HS− combines with H+ near the low-pressure side of the film by the reaction HS− + H+ → H2S; 6. H+ for step (5) is supplied by HCO3− by the reaction HCO3−→ H+ + CO3²−; 7. CO3²− produced in step (6) diffuses back to the high-pressure side;

8. H2S produced in the low-pressure side of the membrane (step 5) leaves the membrane.

Immobilized liquid membranes are prepared by saturating the liquid carrier onto a membrane material. Solid membranes are constructed as hollow-fiber membrane modules to provide a large membrane surface per unit volume and of compact design (Baker, 2012). Solid membranes use two different mechanisms to purify the gases: diffusivity and sorption. A diffusivity difference between two gases indicates that the gas with a smaller molecular size will faster through the porous membrane than the one with larger molecular size (Moioli et al., 2013). Diffusivity is also affected by the shape of the molecules (Faure et al., 2007). The membrane needs to be selected on a number of criteria, including flow rate, operating temperature, feed pressure, permeate pressure, and degree of purification. The main constraints of the membrane separation process are that the membranes are fragile and of high cost. The estimated life time of a membrane for biogas upgrading varies between 5 to 10 years (Bauer et al., 2013). A moisture removal step is generally required to maintain a high permeability and ideal characteristic of a membrane without the need for preconditioning.

11.4 Ex situ removal using sulfur-oxidizing microorganisms

11.4.1 Biological air filtration

Biological gas treatment technology, commonly referred to as biological air filtration, is widely adopted and industrially tested for H2S removal from biogas (Khoshnevisan et al., 2017; Khanongnuch et al., 2019; Estrada et al., 2014; Alvarez-Guzmán et al., 2016). Biofilters (BFs), bioscrubbers (BSs), and biotrickling filters (BTFs) are proven cost-effective and environmentally friendly biological air filtration technologies (Dumont, 2015; Barbusiński et al., 2020). The major differences in these technologies are in their design and operation, however, the mechanism of the air treatment process is similar (Fig. 11.5).

Figure 11.5 Biological air filtration systems for removal of H2S: (A) biofilter; (B) biotrickling filter; and (C) bioscrubber. Adapted and modified from Dumont, 2015, H2S removal from biogas using bioreactors: a review, Int. J. Energy Environ. 6 (2015) 479–498, and Syed et al., 2006, Removal of hydrogen sulfide from gas streams using biological processes—a review, (2006).

BFs consist of a packed bed which is generally composed of peat, compost, and synthetic media with high nutrient availability, high buffer capacity, and moisture retention capacity to ensure a healthy biofilm is formed by the microbiota (Fig. 11.5A) (Syed et al., 2006). In BFs, H2S removal from biogas takes place by the mechanisms of H2S transfer from the gaseous to liquid phase followed by diffusion into the biofilm formed on the inert packing material and biodegradation. In BTFs, a liquid phase containing nutrients stored below a bed of inert packing materials is continuously spread from the top of the inert packing bed containing biofilms (Fig. 11.5B). A liquid circulation in BTF offers several benefits of temperature control, pH control, substrate diffusion to the biofilm, nutrient addition, and removal of metabolites formed during biodegradation. BS (Fig. 11.5C) generally involves a two-stage process where absorption of H2S takes place in the liquid phase present in an absorber unit packed with inert material, followed by the biological treatment in a liquid phase of suspended growth bioreactor. The biological removal of H2S in biological air filters can be achieved under both aerobic and anoxic conditions. Aerobic systems require the addition of a small quantity of air, where O2 acts as an electron acceptor during the biodegradation, whereas under anoxic conditions alternative electron acceptors such as nitrates (NO3−) are utilized. The former system may pose safety risks due to the creation of potentially explosive oxygen/methane mixtures and could lead to biogas dilution due to uncontrolled air addition (Dumont, 2015). In general, reactor configuration (countercurrent and concurrent flow of liquid), operating conditions such as pH of the recirculating liquid, trickling liquid velocity (TLV, m/h) or recirculation rates, empty bed retention time (EBRT, s) and H2S loading rates (LR, gS-H2S/m³/h) influence the elimination capacity (EC, gS-H2S/m³/h) of the biological air filtration systems like BTFs. The application of biological air filtration at both aerobic and anoxic operation

conditions are described in the following sections.

11.4.1.1 Anoxic biological air filters

Anoxic biological air filters overcome the drawbacks of aerobic systems such as safety problems due to the formation of potentially explosive CH4/O2 mixtures, biogas dilution with nitrogen and oxygen mass transfer limitations (Dumont, 2015). In anoxic biological air filters such as BTF, the H2S removal process takes place under anoxic conditions by sulfur-oxidizing nitrate-reducing (SONR) microorganisms such as Thiobacillus denitrificans, which converts H2S to elemental sulfur and sulfate (depending on the nitrate/sulfide ratio) (Eq. 11.9) using nitrate (NO3−) as the electron acceptor (Soreanu et al., 2008b).

(11.9)

More specifically, anoxic H2S oxidation via autotrophic denitrification can proceed according to the stoichiometry of the sulfide oxidation (Eq. 11.10 and 11.11) reported by Mora et al. (2015):

(11.10)

(11.11)

The stoichiometry of the sulfide oxidation given in Eqs. 11.10 and 11.11 is based on the assumption of C5H7O2N as typical biomass composition. Moreover, Mora et al. (2015) reported the oxidation of elemental sulfur to sulfate in the two-step sulfide oxidation associated with denitritation during the respirometric tests. Soreanu et al. (2008a) reported the possible theoretical routes (Eqs. 11.12– 11.16) of sulfide degradation leading to the formation of sulfur, sulfate, and nitrites or nitrogen (N2) at different levels of theoretical estimations of the nitrate demand (g N-NO3− consumed/g H2S removed). Complete denitrification vs. complete H2S oxidation (ratio N/S=1.6)

(11.12)

Complete denitrification vs. partial H2S oxidation (ratio N/S=0.4)

(11.13)

Partial denitrification vs. complete H2S oxidation (ratio N/S=4)

(11.14)

Partial denitrification vs. partial H2S oxidation (ratio N/S=1)

(11.15)

Overall equation:

(11.16)

Thiobacillus denitrificans and Thiomicrospira denitrificans carry out complete denitrification, that is, the reduction of nitrate to nitrogen represented by the Eqs. 11.12–11.13. Few species, such as Thiobacillus thioparus, can reduce nitrates to nitrites (Eqs. 11.14–11.15) (Dumont, 2015). Many studies have demonstrated that controlling the N/S ratio can effectively control the oxidation of H2S to prevent clogging by S and maintain a suitable pH for the successful operation of anoxic BTFs (Soreanu et al., 2008a,b; Cano et al., 2019; Khanongnuch et al., 2018). Cano et al. (2019) reported that a N/S molar ratio below 0.4 does not affect the H2S removal efficiency from the biogas but increases the S production (up to 98.7%). At 1.46 mol N/mol S and 57 gS-H2S m³/h, Montebello et al. (2012) reported 14% elemental sulfur production. Since operation of the anoxic BTFs at higher N/S ratios involves higher nitrates cost, use of nitrified effluent from wastewater treatment plants, or nitrified anaerobic digestate could be an appropriate choice. However, a higher N/S ratio can decrease the pH of the liquid phase which may affect reduction of nitrate to N2, therefore it is necessary to maintain the neutral pH in the circulating liquid. Several studies have reported the range of optimal operating conditions: loading rates in the range of 130–289 gS-H2S/m³/h, EC values ranging from 100–282 gS-H2S/m³/h, EBRT 32–1080 seconds, and TLV of 10–15 m/h (Cano et al., 2019; Fernández et al., 2013; Montebello et al., 2012; Soreanu et al., 2008a).

11.4.1.2 Aerobic biological air filters

Chemotrophs have been used for the aerobic treatment of H2S, where the H2S is oxidized to elemental sulfur under controlled oxygen-limiting conditions, whereas sulfate is formed if excess oxygen is supplied. The biodegradation of H2S occurs according to the reactions presented in Eqs. 11.16 and 11.17:

(11.17)

(11.18)

In some aerobic sulfide removal systems, the oxidation of sulfide to sulfate proceeds in two steps, the first step involves the oxidation of sulfide to elemental sulfur, which can potentially oxidized to sulfite and sulfate in the second step (Buisman et al., 1990). Elemental sulfur and sulfate are the major byproducts of the oxic biotrickling filters (Tomàs et al., 2009). The SO4²−/S ratio produced varies considerably depending on the supplied O2/H2S ratio. With the increasing H2S loading rates the sulfate concentration decreases, thus lowering the SO4² −/S ratio. The major drawback of elemental sulfur formation is the clogging of the reactor bed resulting in a pressure drop (Tomàs et al., 2009). Tomàs et al. (2009) evaluated the technoeconomic feasibility of a full-scale aerobic biotrickling filter for H2S removal treating 80 m³/h biogas. The study reported a maximum elimination capacity of 170 g H2S m³/h with a time of 180 s during 4.5 months of operation on biogas with an average H2S concentration of 3000 ppmv. The study showed the cost/benefits of biotrickling filters (€3.2 per kg H2S treated) over chemical treatment (€5.8 per kg H2S treated) of biogas for H2S removal with a cost saving of about €2.6 per kg H2S treated. The aerobic biological air filtration processes have been successfully demonstrated at the laboratory and pilot scales, while many have been applied at full-scale commercial plants.

11.4.1.3 Commercial applications of biological air filtration systems

Sulfothane (Veolia, ), THIOPAQ O&G (Shell-Paques), THIOPAQ process (Paques, The Netherlands), Biogasclean QSR and MUW (Biogasclean, Denmark), and Biopuric (Biothane, United States) are some of the commercially available processes developed for the removal of high H2S concentrations from biogas or fuel gas. The commercial processes such as Sulfothane and THIOPAQ are aerobic BS systems that apply a two-step process consisting of a caustic scrubber followed by a biological treatment step to recover spent caustic and for elemental sulfur generation (Fortuny et al., 2008). The Biopuric, Biogasclean QSR, and MUW systems are BTFs that work at a low pH, mesophilic temperature, and microaerophilic conditions (Allegue and Hinge, 2014). The systems from Biogasclean, Denmark, have been installed in facilities with H2S concentrations up to 5% and daily sulfur loads up to 5500 kg. In the chemical scrubbing processes, NaOH reacts with H2S (aqueous) to form either sodium hydrosulfide (NaHS) or precipitation of sodium sulfide (Na2S) in the subsequent step. Formation of sodium sulfide (Na2S) occurs as a result of the reaction of NaHS with NaOH limited by the NaOH availability relative to the amount of H2S which is scrubbed into the solution. In commercial processes like Veolia’s Sulfothane, the biogas goes through an up-flow scrubber column, which is fed with alkaline wash water (H2S free) from the top, removing H2S as NaHS and CO2. During the second step, the wash water (with H2S) will enter a slightly aerated bioreactor where aerobic sulfur-oxidizing bacteria oxidize all H2S to produce elemental sulfur and recover alkalinity. The H2S-free wash water is recycled again in the scrubber tank (first step). The overall two-step process is summarized in Eqs. 11.19 and 11.20:

(11.19)

(11.20)

Likewise, another commercial bioscrubber process, THIOPAQ, works under the similar principle of absorption of H2S under slightly alkaline conditions (pH 8– 9) during the first step and sulfide oxidization into elemental sulfur by sulfuroxidizing bacteria in the bioreactor in the second step.

11.4.2 Microalgal removal of H2S

At present, interest in the application of algae for CO2 sequestration is increasing. The microalgal culture using CO2 present in biogas can be promising to upgrade the biogas, concomitantly providing the treatment of an anaerobic effluent. Microalgal biogas purification/scrubbing of H2S utilizes the photosynthetic ability of algae and its symbiosis with sulfur-oxidizing bacteria to utilize the impurities present in biogas (both CO2 and H2S) (Ramaraj and Dussadee, 2015; Muñoz et al., 2015; Meier et al., 2015; Bahr et al., 2014). The dissolved oxygen concentrations in the algal photobioreactors are above saturation due to the photosynthetic activity of the algae. This high concentration of dissolved oxygen facilitates the oxidation of H2S present in the biogas by sulfur-oxidizing bacteria in the scrubber unit (Meier et al., 2018). Bahr et al. (2014) reported the symbiosis of algae and bacteria in the pilot-scale photobioreactors for the simultaneous removal of H2S and CO2. Fig. 11.6 shows the typical algal-based H2S removal process adopted by Bahr et al. (2014) and Meier et al. (2018). Reported biogas residence times vary between 1–2 hours in the absorption column with 100% removal efficiency (Muñoz et al., 2015). The operation of this combined algal–bacterial process occurs at high pH values using alkaliphilic sulfur-oxidizing bacteria and microalge. Bahr et al. (2014) used Spirulina platensis, a cyanobacterium with an optimum growth pH of 9−10 grown in an open pond and bubble column.

Figure 11.6 Schematic process for biogas upgrading by microalgae. Adapted and modified from Meier et al. (2018) Removal of H2S by a continuous microalgae-based photosynthetic biogas upgrading process. Process. Saf. Environ. Prot. 119 (2018) 65–69.

The microalgal H2S removal technology is yet to be demonstrated in full-scale applications and its sustainability requires operation with natural sunlight, use of nutrient from anaerobic digestate or wastewater, and utilization of microalgae biomass for generation of bioenergy (biogas) in the biorefinery concept.

11.5 In situ H2S removal

11.5.1 In situ microaeration

Naturally occurring microorganisms in anaerobic digesters can be employed for removing the H2S present in the biogas. The process is similar to ex situ H2S removal from biogas based on the conversion of H2S to elemental sulfur by a group of specialized microorganisms of the Thiobacillus family. The microorganisms use CO2 from the digester as their carbon source. A small amount of oxygen, 2%–6% of air in biogas depending on the concentration of H2S in the biogas, is added directly in the anaerobic digesters to carry out the following reaction (Eq. 11.21):

(11.21)

Together with the digested sludge, the precipitated elemental sulfur is removed from the digester. The presence of low oxygen concentrations does not negatively influence the activity of the anaerobes as it is consumed quickly inside the digester (Jeníček et al., 2017). Fig. 11.7 shows two strategies for in situ microaeration (Khoshnevisan et al., 2017). In the first strategy (A), the air is added in the recirculation line of the digestate, while in the second strategy (B), the air is directly added in the headspace of the anaerobic digester with the biogas recirculation. The later strategy is commonly adopted for H2S removal from biogas (Khoshnevisan et al., 2017; Muñoz et al., 2015).

Figure 11.7 Possible aeration methods for in situ H2S removal in anaerobic digesters: (A) microaeration of the liquid phase and (B) microaeration of the gaseous phase. Modified from Khoshnevisan et al. (2017) A review on prospects and challenges of biological H2S removal from biogas with focus on biotrickling filtration and microaerobic desulfurization. Biofuel Res. J. 16 (2017) 741–750. https://doi.org/10.18331/BRJ2016.4.4.6.

The air or oxygen dosing rate (0.03 to 218 L O2/L feed), dosing point (liquid phase and in the headspace of the reactors), biogas residence time (2.5– 10 hours), and process temperature (mesophilic and thermophilic) are important considerations for in situ microaeration (Khoshnevisan et al., 2017). In addition, residual oxygen in the biogas must meet the requirements of the biogas utilization technology that will be employed afterwards. Optimal process control and oxygen supply consistent with the H2S concentration in the biogas is important for a higher efficiency of the in situ microaeration. Jeníček et al. (2017) studied microaeration in seven full-scale wastewater treatment plants with anaerobic reactors varying from 1600–30,000 m³ with dosing of air in the recirculation line of the anaerobic reactor. The start-up of the process took 3–12 weeks, which depended on the optimum oxygen dose and adaptation and growth of the sulfide-oxidizing bacteria. The volumetric dose of air varied between 1%– 3% of the biogas production and the air doses depend on the H2S concentration in the biogas. The overall removal efficiency varied between 74%–99%. One of the major disadvantages of dosing the air into the sludge recirculation pipes is the higher requirements of air dose. This strategy is nevertheless advantageous over direct air supply in the headspace as the later strategy can lead to yellowwhitish deposition of elemental sulfur on the digester walls and pipes which can clog the system.

11.5.2 Dosing iron salts/oxides into the digester

Various forms of iron chlorides, phosphates, or oxides (FeCl2, FeCl3, and FeSO4) can be added either directly into the anaerobic digesters or into the

influent (similar strategies are presented in Fig. 11.7 to reduce H2S with the help of Fe²+ or Fe³+). Eq. 11.22 shows the formation of insoluble ferrous sulfide (FeS) from the addition of ferrous salts (FeCl2). The addition of FeCl2 to the reactor mixed liquid is the most regularly practiced for the removal of H2S. Iron (III) salts such as ferric hydroxides [Fe(OH)3] (Eq. 11.23) and ferric chloride (FeCl3) (Eq. 11.24) can also be added for H2S removal. They react with the produced hydrogen sulfide and form insoluble iron sulfide salts. Due to this precipitation, stripping of H2S into the biogas is prevented.

(11.22)

(11.23)

(11.24)

This method can reduce the H2S concentrations in the biogas down to 200– 100 ppmv. However, this method is less effective in attaining low and stable levels of H2S to meet the quality required for vehicle fuel or injection of the gas into the grid (Allegue and Hinge, 2014).

11.6 Combined chemical-biological processes

The chemical processes alone are costly, while the biological processes like biofilteration have some limitations for treating high H2S concentrations due to the low oxidation rate and requirements of certain operating conditions (pH, humidity, and O2 content). Combined chemical-biological processes can combine the ability of chemical and biological processes to treat high concentrations of H2S in the biogas with greater flexibility. Chung et al. (2003) studied a two-step process for H2S gas treatment. The first step is based on absorption with a chemical solution, followed by biological oxidation with Thiobacillus spp. in the second step. Ferric sulfate is used as an oxidizing agent, reducing it to ferrous sulfate, which swiftly reacts with H2S gas and is reduced to ferrous sulfate in the first step (Eq. 11.25).

(11.25)

(11.26)

In the latter stage (Eq. 11.26), the ferrous sulfate solution is circulated to an aerobic bioreactor, where microorganisms oxidize the ferrous iron to ferric iron. The ferric sulfate solution is then recycled into the first stage of the reactor to repeat the cycle. Chung et al. (2003) reported the optimum operating conditions for the biological sulfide oxidation process as pH below 2.3 and 35°C, which resulted in more than 90% Fe³+ formation. In addition, the optimal glucose concentration in the nutrient medium was 0.1% for the optimal growth of T. ferrooxidans 9. Pagella and De Faveri (2000) studied the combined action of a chemicalbiological process for the removal of H2S from biogas. The process was based on a two-step removal, involving H2S absorption in a ferric solution (during which ferric ions are converted to ferrous iron), and biological oxidation of the ferrous ions in the solution back to ferric iron.

11.7 Comparison of H2S removal techniques

A summary of the advantages and disadvantages of the H2S removal technologies is presented in Table 11.1. In the literature, there are discrepancies about the specific capital cost, operation and maintenance (O&M) cost, methane loss, and energy requirements for different technologies at varying treatment capacities. In a comprehensive review of the selection of biogas technologies by Sun et al. (2015), no significant difference in the capital cost of the physicochemical processes were found. However, O&M cost associated with chemical absorption using solvent were comparatively higher. A selection of an appropriate H2S cleaning technology depends on the utilization of the biogas, capital expenditures (CAPEX), operation and maintenance (O&M) costs, and availability of resources (chemicals, parts, and human resources).

Table 11.1

H2S removal process

Advantages and features

Biological

In situ microaeration

Biological air filtration (BF, BS, BTF)

Operation at ambient temperature and pressure, lo

Microalgal technologies

H2S and CO2 could be simultaneously removed,

Absorption

Water scrubbing

Organic solvents (amine)

High efficiency (>99% CH4), low CH4 losses (<0

Absorption with NaOH and FeCl3

Low electricity requirements, low CH4 losses, op

Adsorption

Activated carbon

Metal oxides (iron-impregnated wood chips, iron oxide pellets)

Simple technology, low investment cost, high rem

PSA

Compact technique, tolerant to impurities, efficien

Membrane

Gas–gas; gas–liquid

BF, biofilter; BS, bioscrubber; BTF, biotrickling filters; O&M, operation and maintenance; PSA, pressure swing adsorption.

11.8 Conclusions

Applications of biogas in turbines, internal combustion engines, fuel cells, and grid injection have different tolerance levels (gas quality) for H2S in biogas. Different physicochemical and biological processes have their advantages and disadvantages in of performance, cost effectiveness, flexibility, and complexity of operation. Therefore the selection of technology requires careful examination of the required gas quality to avoid unnecessary treatment costs. Physicochemical processes are much more efficient in of fast start-up time and higher removal efficiencies compared to some biological processes. However, some biological counterparts can serve as low-cost and environmentally benign technological options for H2S removal. Biological technologies such as aerobic biological air filtration (BF, BS, and BTF) have been fully developed at the commercial scale, while H2S removal via anoxic biological air filtration and microalgae technologies is yet to be demonstrated at the full scale. Similarly, the application of membrane processes in biogas upgrading seems promising and needs more research to bridge the knowledge gaps. Nonetheless, future work on integrated physicochemical and biological processes for the minimization of energy requirements, regeneration of chemicals, and recovery and utilization of CO2 will the sustainability of biogas utilization.

References

Adib et al., 1999 Adib F, Bagreev A, Bandosz TJ. Effect of surface characteristics of wood-based activated carbons on adsorption of hydrogen sulfide. J Colloid Interface Sci. 1999;214:407–415 https://doi.org/10.1006/jcis.1999.6200. Allegue and Hinge, 2014 Allegue L.B., Hinge J., Biogas upgrading – evaluation of methods H2S removal, Danish Technological Institute 2014. Available from: https://doi.org/10.1074/m.M110.002766. Alvarez-Guzmán et al., 2016 Alvarez-Guzmán CL, Oceguera-Contreras E, Ornelas-Salas JT, Balderas-Hernández VE, De León-Rodríguez A. Biohydrogen production by the psychrophilic G088 strain using single carbohydrates as substrate. Int J Hydrog Energy. 2016;41:8092–8100 https://doi.org/10.1016/j.ijhydene.2015.11.189. Angelidaki et al., 2018 Angelidaki I, Treu L, Tsapekos P, et al. Biogas upgrading and utilization: current status and perspectives. Biotechnol Adv. 2018;36:452– 466 https://doi.org/10.1016/j.biotechadv.2018.01.011. Augelletti et al., 2017 Augelletti R, Conti M, Annesini MC. Pressure swing adsorption for biogas upgrading A new process configuration for the separation of biomethane and carbon dioxide. J Clean Prod. 2017;140:1390–1398. Awe et al., 2017 Awe, O.W., Zhao, Y., Nzihou, A., Minh, D.P., Lyczko, N., (2017). A review of biogas utilisation, purification and upgrading technologies, waste and biomass valorization. Available from: https://doi.org/10.1007/s12649016-9826-4. Bahr et al., 2014 Bahr M, Díaz I, Dominguez A, González Sánchez A, Muñoz R. Microalgal-biotechnology as a platform for an integral biogas upgrading and nutrient removal from anaerobic effluents. Environ Sci Technol. 2014;48:573– 581 https://doi.org/10.1021/es403596m.

Baker, 2012 Baker RW. Membrane technology and applications John Wiley & Sons 2012. Bandosz, 2002 Bandosz TJ. On the adsorption/oxidation of hydrogen sulfide on activated carbons at ambient temperatures. J Colloid Interface Sci. 2002;246:1– 20. Barbusiński et al., 2020 Barbusiński K, Urbaniec K, Kasperczyk D, Thomas M. Biofilters versus bioscrubbers and biotrickling filters: state-of-the-art biological air treatment. Biofiltration Promis Options Gaseous Fluxes Biotreat. 2020;29–51 https://doi.org/10.1016/b978-0-12-819064-7.00002-9. Bauer et al., 2013 Bauer F, Persson T, Hulteberg C, Tamm D. Biogas upgrading technology overview, comparison and perspectives for the future. Biofuels Bioprod Bioref. 2013;7:499–511 https://doi.org/10.1002/bbb.1423. Buisman et al., 1990 Buisman CJN, Geraats BG, Ijspeert P, Lettinga G. Optimization of sulphur production in a biotechnological sulphide-removing reactor. Biotechnol Bioeng. 1990;35:50–56 https://doi.org/10.1002/bit.260350108. Cano et al., 2019 Cano PI, Brito J, Almenglo F, Ramírez M, Gómez JM, Cantero D. Influence of trickling liquid velocity, low molar ratio of nitrogen/sulfur and gas-liquid flow pattern in anoxic biotrickling filters for biogas desulfurization. Biochem Eng J. 2019;148:205–213 https://doi.org/10.1016/j.bej.2019.05.008. Cebula, 2009 Cebula J. Biogas purification by sorption techniques. Archit Civ Eng Environ. 2009;2:95–103. Chung et al., 2003 Chung YC, Ho KL, Tseng . Hydrogen sulfide gas treatment by a chemical-biological process: chemical absorption and biological oxidation steps. J Environ Sci Heal - Part B Pestic Food Contam Agric Wastes. 2003;38:663–679 https://doi.org/10.1081/PFC-120023522. Dumont, 2015 Dumont E. H2S removal from biogas using bioreactors: a review. Int J Energy Environ. 2015;6:479–498. Estrada et al., 2014 Estrada JM, Dudek A, Munoz R, Quijano G. Fundamental study on gas-liquid mass transfer in a biotrickling filter packed with polyurethane foam. J Chem Technol Biotechnol. 2014;89:1419–1424

https://doi.org/10.1002/jctb.4226. Faure et al., 2007 Faure F, Rousseau B, Lachet V, Ungerer P. Molecular simulation of the solubility and diffusion of carbon dioxide and hydrogen sulfide in polyethylene melts. Fluid Phase Equilib. 2007;261:168–175 https://doi.org/10.1016/J.FLUID.2007.07.032. Fernández et al., 2013 Fernández M, Ramírez M, Pérez RM, Gómez JM, Cantero D. Hydrogen sulphide removal from biogas by an anoxic biotrickling filter packed with Pall rings. Chem Eng J. 2013;225:456–463 https://doi.org/10.1016/j.cej.2013.04.020. Fortuny et al., 2008 Fortuny M, Baeza JA, Gamisans X, et al. Biological sweetening of energy gases mimics in biotrickling filters. Chemosphere. 2008;71:10–17 https://doi.org/10.1016/j.chemosphere.2007.10.072. Gutiérrez-Padilla et al., 2010 Gutiérrez-Padilla MGD, Bielefeldt A, Ovtchinnikov S, Hernandez M, Silverstein J. Biogenic sulfuric acid attack on different types of commercially produced concrete sewer pipes. Cem Concr Res. 2010;40:293–301 https://doi.org/10.1016/j.cemconres.2009.10.002. Haycarb, 2019 Haycarb, P.L.C., Benefits of Coconut Shell Activated Carbon (2019). https://www.haycarb.com/activated-carbon (accessed November 4, 2019). Jeníček et al., 2017 Jeníček P, Horejš J, Pokorná-Krayzelová L, Bindzar J, Bartáček J. Simple biogas desulfurization by microaeration – full scale experience. Anaerobe. 2017;46:41–45 https://doi.org/10.1016/j.anaerobe.2017.01.002. Kaliva and Smith, 1983 Kaliva AN, Smith JW. Oxidation of low concentrations of hydrogen sulfide by air on a fixed activated carbon bed. Can J Chem Eng. 1983;61:208–212. Kameyama et al., 1983 Kameyama T, Dokiya M, Fujishige M, Yokokawa H, Fukuda K. Production of hydrogen from hydrogen sulfide by means of selective diffusion membranes. Int J Hydrog Energy. 1983;8:5–13. Khanongnuch et al., 2018 Khanongnuch R, Di Capua F, Lakaniemi AM, Rene ER, Lens PNL. Effect of N/S ratio on anoxic thiosulfate oxidation in a fluidized

bed reactor: experimental and artificial neural network model analysis. Process Biochem. 2018;68:171–181 https://doi.org/10.1016/j.procbio.2018.02.018. Khanongnuch et al., 2019 Khanongnuch R, Di Capua F, Lakaniemi A, Rene ER, Lens PNL. Transient – state operation of an anoxic biotrickling filter for H2S removal. J Hazard Mater. 2019;377:42–51 https://doi.org/10.1016/j.jhazmat.2019.05.043. Khoshnevisan et al., 2017 Khoshnevisan B, Tsapekos P, Alfaro N, et al. A review on prospects and challenges of biological H2S removal from biogas with focus on biotrickling filtration and microaerobic desulfurization. Biofuel Res J. 2017;16:741–750 https://doi.org/10.18331/BRJ2017.4.4.6. Krich et al., 2005 Krich K, Augenstein D, Batmale JP, Benemann J, Rutledge B, Salour D. Biomethane from dairy waste. A Sourceb Prod Use Renew Nat Gas Calif. 2005;147–162. Magomnang and Villanueva, 2014 Magomnang A., Villanueva E.P., Removal of hydrogen sulfide from biogas using dry desulfurization systems, in: Int. Conf. Agric. Environ. Biol. Sci. Phuket (Thailand), DOI, 2014. Martin, 2008 Martin II JH. A New Method to Evaluate Hydrogen Sulfide Removal from Biogas North Carolina State University 2008; https://repository.lib.ncsu.edu/bitstream/handle/1840.16/1047/etd.pdf? sequence=1&isAllowed=y. Matson et al., 1977 Matson SL, Herrick CS, Ward WJ. Progress on the selective removal of H2S from gasified coal using an immobilized liquid membrane. Ind Eng Chem Process Des Dev. 1977;16:370–374 https://doi.org/10.1021/i260063a022. Medrano et al., 2019 Medrano JA, Llosa-Tanco MA, Tanaka DAP, Gallucci F. Membranes utilization for biogas upgrading to synthetic natural gas. Substit Nat Gas from Waste Elsevier 2019;245–274. Meier et al., 2015 Meier L, Pérez R, Azócar L, Rivas M, Jeison D. Photosynthetic CO2 uptake by microalgae: an attractive tool for biogas upgrading. Biomass Bioenergy. 2015;73:102–109 https://doi.org/10.1016/j.biombioe.2014.10.032.

Meier et al., 2018 Meier L, Stara D, Bartacek J, Jeison D. Removal of H2S by a continuous microalgae-based photosynthetic biogas upgrading process. Process Saf Environ Prot. 2018;119:65–69 https://doi.org/10.1016/j.psep.2018.07.014. Moioli et al., 2013 Moioli S, Pellegrini LA, Picutti B, Vergani P. Improved ratebased modeling of H2S and CO2 removal by methyldiethanolamine scrubbing. Ind Eng Chem Res. 2013;52:2056–2065 https://doi.org/10.1021/ie301967t. Montebello et al., 2012 Montebello AM, Fernández M, Almenglo F, et al. Simultaneous methylmercaptan and hydrogen sulfide removal in the desulfurization of biogas in aerobic and anoxic biotrickling filters. Chem Eng J 200–. 2012;202:237–246 https://doi.org/10.1016/j.cej.2012.06.043. Mora et al., 2015 Mora M, Fernández M, Manuel Gómez J, et al. Kinetic and stoichiometric characterization of anoxic sulfide oxidation by SO-NR mixed cultures from anoxic biotrickling filters. Appl Microbiol Biotechnol. 2015;99:77–87 https://doi.org/10.1007/s00253-014-5688-5. Muñoz et al., 2015 Muñoz R, Meier L, Diaz I, Jeison D. A review on the stateof-the-art of physical/chemical and biological technologies for biogas upgrading. Rev Environ Sci Biotechnol. 2015;14:727–759 https://doi.org/10.1007/s11157015-9379-1. Ortiz et al., 2014 Ortiz FJG, Aguilera PG, Ollero P. Biogas desulfurization by adsorption on thermally treated sewage-sludge. Sep Purif Technol. 2014;123:200–213. OSHA, 2020 OSHA, Safety and Health Topics | Hydrogen Sulfide - Hazards | Occupational Safety and Health istration (n.d.). https://www.osha.gov/SLTC/hydrogensulfide/hazards.html (accessed July 2, 2020). Pagella and De Faveri, 2000 Pagella C, De Faveri DM. H2S gas treatment by iron bioprocess. Chem Eng Sci. 2000;55:2185–2194 https://doi.org/10.1016/S0009-2509(99)00482-0. Petersson and Wellinger, 2009 Petersson A, Wellinger A. Biogas upgrading technologies–developments and innovations. IEA Bioenergy. 2009;20:1–19. Qi and Cussler, 1985 Qi Z, Cussler EL. Hollow fiber gas membranes. AIChE J.

1985;31:1548–1553. Ramaraj and Dussadee, 2015 Ramaraj R, Dussadee N. Biological purification processes for biogas using algae cultures: a review. Int J Sustain Green Energy. 2015;4:20–32 https://doi.org/10.11648/j.ijrse.s.2015040101.14. Rasi et al., 2007 Rasi S, Veijanen A, Rintala J. Trace compounds of biogas from different biogas production plants. Energy. 2007;32:1375–1380 https://doi.org/10.1016/j.energy.2006.10.018. Rezakazemi et al., 2017 Rezakazemi M, Heydari I, Zhang Z. Hybrid systems: combining membrane and absorption technologies leads to more efficient acid gases (CO2 and H2S) removal from natural gas. J CO2 Util. 2017;18:362–369 https://doi.org/10.1016/j.jcou.2017.02.006. Ryckebosch et al., 2011 Ryckebosch E, Drouillon M, Vervaeren H. Techniques for transformation of biogas to biomethane. Biomass Bioenergy. 2011;35:1633– 1645 https://doi.org/10.1016/J.BIOMBIOE.2011.02.033. Salas et al., 2012 Salas B, Wiener MS, Badilla GL, et al. H2S pollution and its effect on corrosion of electronic components. In: Badilla M, Valdez GL, Schorr B, eds. Air Quality – New Perspective. INTECHOPEN 2012;263–286. https://doi.org/10.5772/39247. San Miguel et al., 2001 San Miguel G, Lambert SD, Graham NJD. The regeneration of field-spent granular-activated carbons. Water Res. 2001;35:2740–2748. Shareefdeen et al., 2005 Shareefdeen Z, Herner B, Singh A. Biotechnology for air pollution control—an overview. Biotechnol Odor Air Pollut Control Springer 2005;3–15. Soreanu et al., 2008a Soreanu G, Béland M, Falletta P, Edmonson K, Seto P. Investigation on the use of nitrified wastewater for the steady-state operation of a biotrickling filter for the removal of hydrogen sulphide in biogas. J Environ Eng Sci. 2008a;7:543–552 https://doi.org/10.1139/S08-023. Soreanu et al., 2008b Soreanu G, Béland M, Falletta P, Edmonson K, Seto P. Laboratory pilot scale study for H2S removal from biogas in an anoxic biotrickling filter. Water Sci Technol. 2008b;57:201–207

https://doi.org/10.2166/wst.2008.023. Sun et al., 2015 Sun Q, Li H, Yan J, Liu L, Yu Z, Yu X. Selection of appropriate biogas upgrading technology—a review of biogas cleaning, upgrading and utilisation. Renew Sustain Energy Rev. 2015;51:521–532 https://doi.org/10.1016/j.rser.2015.06.029. Syed et al., 2006 Syed M., Soreanu G., Falletta P., Béland M., Removal of hydrogen sulfide from gas streams using biological processes—a review, (2006). Thrän et al., 2014 Thrän, D., Billig E., Persson T., Svensson M., Daniel-Gromke J., Ponitka J., et al., Biomethane-Status and Factors Affecting Market Development and Trade, 2014. http://task40.ieabioenergy.com/wpcontent/s/2013/09/t40-t37-biomethane2014.pdf. Tomàs et al., 2009 Tomàs M, Fortuny M, Lao C, Gabriel D, Lafuente J, Gamisans X. Technical and economical study of a full-scale biotrickling filter for H2S removal from biogas. Water Pract Technol. 2009;4 https://doi.org/10.2166/wpt.2009.026. Wellinger and Lindberg, 2007 Wellinger A., Lindberg A., Biogas Upgrading and Utilization, 2007. http://www.iea-biogas.net/files/datenredaktion//publi-task37/Biogasupgrading.pdf. Zheng et al., 2014 Zheng Y, Brown B, Nešić S. Electrochemical study and modeling of H2S corrosion of mild steel. Corrosion. 2014;70:351–365 https://doi.org/10.5006/0937. Zicari, 2003 Zicari S.M., Removal of hydrogen sulfide from biogas using cow manure compost, Cornell University, 2003. Zulkefli et al., 2016 Zulkefli NN, Masdar MS, Jahim J, Majlan EH. Overview of H2S removal technologies from biogas production. Int J Appl Eng Res. 2016;11:10060–10066.

Chapter 12

Biological upgrading of biogas through CO2 conversion to CH4

Michael Vedel Wegener Kofoed, Mads Borgbjerg Jensen and Lars Ditlev Mørck Ottosen, Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark

Abstract

The societal transition from one based on energy from fossil sources to one based on renewable energy requires the development of new methods for both energy production and storage. Biogas is today upgraded to biomethane of natural gas quality by conventional gas scrubbing technologies. However, these conventional technologies release the CO2 from biogas to the atmosphere without any further valorization. We here present the concept of biological methanation (or biomethanation), as an alternative technology which can be used for (1) biogas upgrading to natural gas quality, (2) conversion of electrical energy for subsequent storage, and (3) CO2 capture and utilization. Biomethanation is a power-to-gas technology which utilizes microorganisms of the domain Archaea as catalysts, and knowledge about the microorganisms and the factors controlling their activity is therefore essential to understand and further develop the technology. This chapter introduces some of the basic concepts of biomethanation along with recent developments within both academia and industry. The further development and implementation of the biomethanation technology is moreover evaluated in the context of the drivers and challenges controlling the transition to a world based on renewables.

Keywords

Renewable energy; biological methanation; biomethanation; power-to-gas technology; CO2 conversion; hydrogenotrophic methanogenesis; anaerobic microbiology; reactor designs; gas–liquid mass transfer; CSTR; trickling bed reactors

Chapter outline

Outline

12.1 Biogas upgrading 321

12.2 Hydrogen generation and utilization 324

12.3 Methanation 326

12.4 Microbial basis for biomethanation 327

12.4.1 Methanogens 327 12.4.2 Processes in anaerobic digestion 328

12.5 Reactor configurations 330

12.5.1 In situ biomethanation 331 12.5.2 Ex situ biomethanation 332

12.6 Factors controlling biomethanation 337

12.6.1 Mass transfer of H2337 12.6.2 Temperature 343 12.6.3 Growth requirements 344 12.6.4 pH and CO2346 12.6.5 Bacterial interaction and competition 348

12.7 Reactor design for biological methanation 350

12.7.1 Continuous stirred tank reactor 350 12.7.2 Trickle-bed reactors 352

12.8 Future perspectives and applications 354


Abbreviations list 357

References 357

12.1 Biogas upgrading

Biogas produced from biological anaerobic degradation of organic matter has a high concentration of both methane (CH4) and carbon dioxide (CO2). Biogas has classically been utilized for heat or combined heat and power production using gas engines, which only requires removal of corrosive trace contaminants like hydrogen sulfide (H2S). Heat and power production remains a key application of biogas; but increased focus on substituting fuels of fossil origin, as well as breakthrough of cost-effective wind and solar power production, changes the financial incentives in the biogas sector leading to increased implementation of biogas upgrading technologies for production of purified CH4 as a renewable natural gas substitute (Angelidaki et al., 2018; Scarlat et al., 2018). Except for CO2, most of the gas impurities are present in relatively low concentrations in raw biogas. The concentration of CO2 in the raw biogas depends on the organic feedstock used in the anaerobic digester, but is typically 40%–50% (Weiland, 2010). The high concentrations of CO2 reduce the calorific value and the Wobbe index, which are measures for the heating value of biogas, and consequently make raw biogas unsuitable for direct substitution of natural gas. To reach natural gas quality, the raw biogas has to be purified of minor contaminants such as oxygen (O2), nitrogen (N2), hydrogen (H2), and H2S, and the CO2 has to be either removed or converted to CH4 (Table 12.1; Sun et al., 2015). Classical upgrading techniques include physical or chemical removal of the CO2 without further valorization and are described in other chapters of this book. The separation techniques for biogas upgrading release the separated CO2 to the atmosphere, where it contributes to the greenhouse effect.

Table 12.1

EU

US

Min. CH4 concentration

70–98 % mol

93.5–95.5 % mol

Max. CO2 concentration

1–8 % mol

2–3 % mol

Max. O2 concentration

0.01–1 % mol

0.2–3 % mol

Max. N2 concentration

2–10 % mol

Max. H2 concentration

0.1 % mol

0.1 % mol

Max. H2S concentration

2–15 mg/m³

6–88 mg/m³

Max total sulfur concentration

10–150 mg/m³

265 mg/m³

Water content

0.05–1 g/m³

65 mg/m³

The upgraded and purified biogas (termed biomethane or biosynthetic natural gas, bioSNG) can directly replace fossil-based natural gas and therefore be used for various applications including: (1) injection and bulk storage in the existing natural gas grid; (2) fuel in domestic stoves and boilers; and (3) carbon-based biofuel for the transport sector (Sun et al., 2015). In recent years, there has been an increasing interest in capturing and reusing CO2, both directly for industrial purposes, but also as a chemical building block for the formation of renewable fuels and commodity chemicals, including methanol, syngas, alkanes, or methane (Al-Mamoori et al., 2017). Methanation is part of a power-to-gas process chain, which can be used for valorization of waste gas CO2 by capture and conversion to CH4 through the reaction with H2 [Eq. (12.1)] (Götz et al., 2016). Methanation, as illustrated in Fig. 12.1, thus constitutes a technology platform for biogas upgrading, where the CO2 in biogas is converted to CH4, rather than emitted to the atmosphere, as is otherwise done by conventional biogas upgrading technologies based on gas separation (Angelidaki et al., 2018).

Figure 12.1 Production of CH4/synthetic natural gas/biomethane based on power-to-gas conversion of renewable electricity and CO2. Methanation can be catalyzed either chemically or biologically.

Methanation can be catalyzed either chemically using a metal catalyst (Sabatier process) or biologically through the use of anaerobic microorganisms. This chapter focuses on biological methanation technologies for upgrading of biogas to natural gas quality. Although the catalysts of the chemical and biological methanation differ, they share the same reaction scheme [Eq. (12.1)], requiring the addition of four moles of H2 to reduce one mole of CO2 into one mole CH4 with two moles of water formed as a byproduct:

(12.1)

The reaction stoichiometry reveals that significant reducing equivalents are needed to reduce the one-carbon atom from its most oxidized form (CO2; oxidation state +4) to its most reduced form (CH4; oxidation state −4). Any source of CO2 can in principle be captured and converted to CH4 via methanation, but biogas constitutes a highly suitable CO2 source for the process due to its high concentration of CO2, relatively low concentration of other contaminants like N2 and O2, and existing technical capacity at biogas plants to handle biomethane. Simultaneously with biogas upgrading through CO2 conversion to CH4, methanation enables transformation of electricity to CH4 by using H2 generated from electrolysis. It is important to note here that the use of electricity for biogas upgrading is only sound if the input electricity is derived from renewable sources. Methanation-based biogas upgrading is therefore developed in the context of a societal transition shifting from a fossil-based energy system to that of a renewable-based one, where power-to-gas technologies enable energy production and consumption to be balanced (IRENA, 2019a).

12.2 Hydrogen generation and utilization

The current way of generating H2 at industrial scale is incompatible with sustainable methanation of CO2, as more than 95% of global H2 production is based on fossil fuels (IRENA, 2018b), with the majority originating from steamreforming of fossil CH4 [Eq. (12.2)].

(12.2)

H2 produced by splitting water into its elemental constituents using renewable electricity through the process of electrolysis [Eq. (12.3)] constitutes an alternative to the fossil-based production of H2.

(12.3)

Water electrolysis can be based on different technologies including solid oxide electrolysis (SOEC), alkaline electrolysis, and polymer electrolyte membrane electrolysis. Although the latter two technologies have been developed to a commercial scale, only about 4% of global H2 production is currently based on electrolysis (Götz et al., 2016; IRENA, 2018b). Methanation technologies are therefore completely reliant on the continued development and implementation of water electrolysis systems as well as the continuous expansion of global renewable power production, for providing an alternative and sustainable source of H2. This development is ed by societal and political ambitions to

reduce the anthropogenic CO2 footprint through increasing production of renewable energy. With the development and increased focus on energy generation from renewable sources using wind turbines and photovoltaic solar s, the formation of H2 through electrolysis is here recognized as a key player in the future energy system based on renewables (IRENA, 2018b). There are several societal drivers ing the development and implementation of both electrolysis and methanation: (1) The need to find sustainable renewable fuel alternatives to fossil fuels currently employed in the transportation sector (Mathiesen et al., 2015). (2) The development and increased implementation of wind turbines and photovoltaic technology, which have reduced the prices of renewable electricity (IRENA, 2019b), and hereby increased the feasibility of producing H2 sustainably. (3) In countries with increasing reliance on electricity from fluctuating renewable power sources such as wind and solar, a new demand for electricity storage has arisen to balance production and demand. Due to the fluctuating nature of wind and sun, stable electricity production from these renewable sources alone is challenging, and gives rise to high price volatility based on imbalanced supply and demand (Lewandowska-Bernat and Desideri, 2017; Sovacool, 2013). Fig. 12.2 shows the fluctuations in electricity production from Denmark (“Nord Pool”), which had 47% of the demand covered by wind in 2019 (Gronholt-Pedersen, 2020). The fluctuations illustrate the need for technologies enabling large-scale electricity storage in periods where production exceeds demand to also be able to generate electricity in periods of low electricity production. As a chemical energy carrier, H2 has potential as a transport fuel and to balance fluctuating electricity production and demand, but the lack of existing infrastructure to handle H2 gives it limited application at present (Singh et al., 2015). Instead, methanationbased CH4 produced from H2 [Eq. (12.1)] offers a versatile option for immediate large-scale production, application, and storage of an energy-rich molecule. CH4 is hence directly applicable as both a transport fuel and for largescale balancing of the electricity system due to its compatibility with the existing natural gas infrastructure, including natural gas vehicles, the natural gas grid, and natural gas storage reservoirs (Fig. 12.1). Comparing the net calorific value of H2 (ΔHcΘ=241.82 kJ/mol) versus that of CH4 (ΔHcΘ=802.62 kJ/mol), it is clear that the conversion of four moles of H2 to one mole of CH4 results in an energy loss of ~17% (standard heat of combustion values are sourced from Poling et al., 2008). However, the existence of a comprehensive natural gas infrastructure for easy CH4 storage and transportation in many ways compensate for the energy loss that is inherent to the conversion from H2 to CH4 (Götz et

al., 2016). The superior storage capacity is exemplified by comparing energy storage capacity in the Danish national gas grid (11,093 GWh) with that of the Hornsdale Power Reserve (0.129 GWh) in Australia, which is currently the largest battery in operation (“//hornsdalepowerreserve.com.au/,” 2019). Although the conversion of H2 and CO2 to CH4 is associated with some loss, the production of high-quality CH4 through methanation of renewable H2 enables large-scale energy storage in regions with a natural gas infrastructure at almost no loss during long-term storage, and at a price and capacity that are clearly superior to other storage technologies like batteries (Lund et al., 2016). Although comparisons of energy storage technologies are subject to constant change as new technologies, especially within battery technology, are rapidly progressing, Lund et al. (2016) estimated that the investment cost of large-scale sodium batteries was 600,000 euro/MWh, whereas the investment cost for a gas cavern for CH4 storage was 46 euro/MWh (Lund et al., 2016). However, the exact usage and market breakthrough of CH4 produced via methanation technologies as transport fuel or for large-scale power balancing will depend not only on the direct investment cost of the storage technology but also on the existing regional energy infrastructure.

Figure 12.2 Examples of fluctuations in renewable electricity production in Denmark, which is a country where wind electricity is continuously expanding. The Danish electricity consumption was more than covered by wind-generated electricity throughout a 24 h period (September 15, 2019), while hardly any renewable electricity was produced on September 18, 2017. Data source: Nord Pool [WWW Document], <www.nordpoolgroup.com>.

12.3 Methanation

The most mature methanation technology for biogas upgrading is chemical methanation according to the Sabatier process [Eq. (12.1)]. Chemical methanation is operated at high temperature (~300°C) and high pressure (5−20 MPa) in order to secure purposeful reaction kinetics and conversion rates, and hereby minimize capital and operational expenditures. For biogas upgrading, the chemical catalyst utilized for chemical methanation has a high sensitivity to impurities in raw biogas, where even minute concentrations of H2S are a primary concern due to its poisonous effect on the metal catalyst. Upstream purification of biogas is therefore absolutely essential in chemical methanation (Benjaminsson et al., 2013). Current activities within the field of chemical methanation aim at increasing the economy of the technology by lowering operating expenses, for example, by using the reaction heat from the exothermic process of methanation for heat-consuming H2 production in SOEC (Dannesboe et al., 2020). Chemical methanation is at a high technological readiness level and currently is deployed at a pilot scale at different sites (Thema et al., 2019). An alternative to chemical methanation is biological methanation technology catalyzed by microorganisms. There has been substantial research on biological methanation (or biomethanation) in the past decade, because it can be operated at atmospheric pressure and moderate temperatures, and its biological catalyst is self-renewing and much more tolerant to impurities compared to the chemical catalysts. The catalysts in biological methanation are methane-producing microorganisms (methanogens), which catalyze the reduction of CO2 to CH4 using H2 as an electron donor to gain energy for cellular maintenance and biomass growth (Zabranska and Pokorna, 2018). Due to the lower temperature and pressure of operation compared to chemical methanation, biomethanation has a lower CH4 production rate than the chemical process (Götz et al., 2016). However, its milder process pressure and temperature conditions result in significantly lower requirements for the process equipment. The higher robustness to gas impurities additionally implies a reduced need for gas purification in biological methanation compared to chemical methanation. Based on the above, it has been estimated that biological methanation can prove a cost-

efficient technology regarding capital expenditures and to a lesser degree with regards to operational cost (Benjaminsson et al., 2013). However, the maturity of biomethanation technology is currently at a significantly lower level than chemical methanation, which makes it difficult to reliably estimate exact capital and operational expenditures. Further development of reactor and process designs at an industrial scale are expected along with increased knowledge of the biological basis for biomethanation. Although biological methanation is not as mature a technology as its chemical counterpart, it poses a promising alternative or supplement to the chemical process.

12.4 Microbial basis for biomethanation

12.4.1 Methanogens

Biological methanation is catalyzed by hydrogenotrophic methanogens, which are microorganisms of the domain Archaea. The methanogens use the energy gained from the reduction of CO2 with H2 for cellular maintenance and growth and are hence capable of replicating and repairing themselves. Methanogens can be divided into three groups based on their ability to use different substrates (Holmes and Smith, 2016):

1. Hydrogenotrophic methanogens which can utilize H2 and CO2 (and in some cases formate and alcohols) for the formation of CH4. 2. Methylotrophic methanogens which can utilize methylated substrates like methanol, methylamines, and methyl sulfides for the reduction of CO2 to CH4. 3. Acetoclastic methanogens which produce CH4 (and CO2) from acetate.

Pure cultures of methanogens have been isolated and characterized from a range of different natural and technical systems, including anaerobic digesters. The majority of the isolates have been found to be able to utilize H2 and CO2 for the formation of CH4 and the ability to metabolize H2 and CO2 is an omnipresent and critical trait of all anaerobic digesters. In fact, H2 and CO2 are end-products from the fermentation of organic material that play an integral part in biogas production through anaerobic digestion. Digestates from anaerobic digesters are consequently an often used inoculum for biomethanation reactors, and the biomethanation reactors therefore often include some—or all—of the process

steps known from conventional anaerobic digesters (see Section 12.5). Understanding the processes of anaerobic digestion is therefore essential to understand the processes involved in biomethanation, as well as those interacting with it.

12.4.2 Processes in anaerobic digestion

Biogas production through anaerobic digestion is in essence a biological degradation and gasification process, where organic matter is degraded to gaseous end-products in the form of CO2 and CH4. Anaerobic digesters handle a wide range of different organic feedstocks, including agricultural and industrial waste, municipal organic waste, along with waste products from wastewater treatment (Weiland, 2010). The composition of these waste products is highly complex and consists of a range of polymers based on sugars, amino acids, and fatty acids. During biogas production, these organic polymers are degraded to CH4 and CO2 through an intricate web of microbial processes, catalyzed by a highly diverse community of microorganisms. The microbial community is formed by several factors like temperature and ammonia concentrations (De Vrieze et al., 2015), which can affect both the number and identity of the microbial species in the biogas reactor systems (Levén et al., 2007). Regardless of the operational conditions and complexity of the anaerobic digestion process, it can be summarized in four major steps: hydrolysis, acidogenesis, acetogenesis, and methanogenesis (Fig. 12.3).

Figure 12.3 Schematic of the four overall microbial steps in anaerobic digestion.

During the first step of anaerobic biomass degradation, hydrolytic bacteria are responsible for enzymatic hydrolysis of the various types of organic polymers into soluble monomers and oligomers like sugars, fatty acids, and amino acids. Acidogenic microorganisms subsequently degrade the soluble hydrolysis products to smaller organic compounds like short-chain volatile fatty acids (VFAs), alcohols, acetate, H2, and CO2. The primary fermentation by acidogenesis is followed by a secondary step of fermentation termed acetogenesis. Here acetogenic bacteria oxidize the VFAs and alcohols produced via acidogenesis to acetate, CO2, and H2. In the final step of anaerobic digestion, CH4 is produced, mainly from acetate or H2 and CO2, by acetoclastic and hydrogenotrophic methanogenesis, respectively. Acetate can also be formed from H2 and CO2 through the processes of homoacetogenesis [Eq. (12.4)] or degraded to H2 and CO2 by acetate-oxidizing bacteria.

(12.4)

Homoacetogenesis is of particular interest in biomethanation systems because the substrates for this process are similar to those of hydrogenotrophic methanogenesis [Eq. (12.1)]. These two reactions are therefore competing and their balance is essential for CH4 yield and reactor stability, as described in Section 12.6.5. The acetate produced from homoacetogenesis does not necessarily constitute a loss of H2 and CO2 in the process, as acetate can be converted to CH4 either directly through the action of acetoclastic methanogens, or in a two-step process where syntrophic acetate-oxidizers (SAOs) convert the acetate to H2 and CO2, which can then be converted further to CH4 by the hydrogenotrophic methanogens. The SAOs are sensitive to high concentrations of H2 and their activity at high concentrations of H2 is therefore thought to be limited.

Some of the other acetogenic reactions which produce H2 as part of the fermentation processes are also of interest in relation to biomethanation fueled by the addition of H2 from an external source (exogenous H2). The H2producing reactions of acetogenesis are endergonic even at low concentrations of H2 and are only made thermodynamically favorable due to the H2-consuming processes, not least methanogenesis, that keep the H2 concentrations sufficiently low (Table 12.2). Due to their strong interdependency, the H2 producers and consumers are often found to be tightly coupled in a syntrophic relationship to facilitate the transfer of H2 directly from producer to consumer (Madigan et al., 2019a). In situations where H2-producing and -consuming reactions are imbalanced and H2 concentrations increase, VFAs accumulate in the digestate, which first the lowers biogas yield but may ultimately be fatal to the stability of anaerobic digesters in the case of excess acidification (Weiland, 2010). The addition of exogenous H2 in the context of power-to-gas can for the same reason inhibit acetogenic processes and result in accumulation of VFAs like butyric and propionic acid (Table 12.2). The potential accumulation of VFAs in response to H2 addition to anaerobic digesters has been the subject of several studies and is discussed in more detail in Section 12.5.1.

Table 12.2

Free energy change Reactions

Free energy change ΔG ’ (kJ/reaction)a

Glucose + 4H2O → 2 acetate− + 2HCO3− + 4H+ + 4H2

−207

Butyrate− + 2H2O → 2 acetate− +H+ + 2H2

+48.2

Propionate− + 3H2O → acetate− + HCO3− + H+ + H2

+76.2

2 Ethanol + 2H2O → 2 acetate− + 2H+ + 4H2

+19.4

aStandard conditions, solutes 1 M gases, 1 atm., 25°C.

Many of the reported studies of biomethanation have employed digestate from anaerobic digesters as the microbiological basis for their setup (e.g., Dupnock and Deshusses, 2017; Luo and Angelidaki, 2012; Strübing et al., 2017). The use of this complex inoculum for biomethanation has often been chosen due to the lower cost of the microbial inoculum, the nutrients needed and requirements for process sterility, compared to the use of pure culture methanogens. Biomethanation reactors based on digestate inocula are inherently inhabited by a complex community of microbes, but the relative abundancy of hydrogenotrophic methanogens increases due to the continued addition of their growth-limiting substrate: H2 (Agneessens et al., 2017; Jensen et al., 2019). An alternative to the use of complex microbial cultures is the use of pure axenic cultures of hydrogenotrophic methanogens (Martin et al., 2013; Seifert et al., 2014). The methanogenic processes are here the same but without the interacting and competing processes of the complex inoculum. Although the use of pure cultures allows for simpler process control, it demands a high degree of process hygiene to maintain it as a pure culture (Dupnock and Deshusses, 2019). Use of pure cultures would therefore require sterilization not only of liquids but also of the added biogas, thereby increasing the investment and operational costs of the biomethanation system.

12.5 Reactor configurations

The application of both pure and complex cultures for biomethanation has been tested in different reactor configurations and setups which can be categorized in two main groups: in situ or ex situ. The refer to whether the process of methanation of electrolysis-derived H2 takes place directly within the anaerobic digester alongside the complex network of microbial processes depicted in Fig. 12.3 (in situ), or in a dedicated methanation reactor, separated from the main anaerobic digester (ex situ) (Fig. 12.4). The application of the in situ configuration is constrained to biogas upgrading, while the ex situ reactor can be used for methanation of any CO2 source compatible with methanogenic activity, including raw biogas and CO2 purified from conventional biogas upgrading technologies. Both in situ and ex situ concepts are currently being developed both in academia and industry. The simultaneous development of in situ and ex situ methanation platforms reflects the general low technology readiness level of biomethanation, with the main development still taking place in lab-scale reactors. Efforts to upscale the biomethanation technology are outlined in Section 12.7.

Figure 12.4 In situ and ex situ reactor concepts for biomethanation. (A) In the in situ configuration, the biological processes part of organic matter degradation form the CO2 needed for methanation. (B) In the ex situ configuration, CO2 can be supplied both as purified CO2, and as raw biogas, which is a mixture of CO2 and CH4.

12.5.1 In situ biomethanation

The in situ biomethanation process takes place within the primary anaerobic digester, which is designed to handle the conventional biogas production based on the anaerobic degradation of organic matter (Fig. 12.4). In situ biomethanation thus proceeds through H2 addition to the anaerobic digester, where the digester’s naturally residing microbial population catalyzes the methanation reaction simultaneously with organic matter degradation. A key value proposition of in situ methanation is to reduce the cost for constructing and operating a separate reactor by utilizing existing reactor volume for biomethanation of the externally supplied H2. The internal supply of CO2 from biomass degradation means that only H2 has to be added to the process. A critical point in in situ biomethanation is the addition of exogenous H2, which is related to high operational costs and subject to intense research (see Section 12.6), because H2 is poorly soluble, yet must solubilize to be converted by the methanogens. H2 addition is ideally combined with the agitation normally used for keeping the biomass in suspension to lower the operational costs of the methanation process (Jensen et al., 2018, 2020a). While in situ methanation entails the use of existing reactor capacity, it is essential that the added H2 does not interfere with other anaerobic processes and that the anaerobic digester maintains its primary function: biological degradation and gasification of the organic waste material. When implementing in situ methanation it is therefore important to ensure stable and efficient biogas production in parallel with the methanation-based biogas upgrading process. The H2-sensitive processes exemplified in Table 12.2 intrinsic to anaerobic organic matter degradation have consequently been the focus of several biomethanation

research studies addressing the concern that addition of exogenous H2 would inhibit acetogenesis. However, most studies find that only limited accumulation of VFAs occurs following H2 addition, which has been hypothesized to be due to a combination of H2’s low solubility combined with its fast consumption by the H2-consuming microorganisms (Agneessens et al., 2018). In a study on microscale gradients of H2 in anaerobic slurry, Garcia-Robledo et al. (2016) found that H2 was consumed within a zone of <0.5 mm adjacent to the gas– liquid interface, when supplying 100% H2 to digestate from an anaerobic digester, which had not previously been subjected to exogenous H2 (GarciaRobledo et al., 2016). In the study, the digestate was mixed with glass beads to ensure a stable zone for microsensor profiling, thereby reducing the number of methanogens per volume. The zone of consumption is therefore expected to be even narrower in an adapted and active biomethanation community, where the methanogens will be more abundant. These results suggest that the added H2 only reaches a minor part of the microorganisms under realistic reactor operation. Acetogenesis will thus only be inhibited very locally in the proximity of an H2 bubble, while proceeding undisturbed outside this zone, explaining the common observations of only limited VFA accumulation during in situ biomethanation. Although low penetration of H2 from the gas phase into the liquid reduces the challenges of process inhibition, it also poses a significant challenge for distributing H2 to all parts of the reactor where CO2 is being continuously produced. Failure to do so will result in the release of unconverted CO2 and H2, and only result in partial biogas upgrading. The production of natural gas-quality methane through in situ biomethanation has until now only been validated in the lab scale (Table 12.5), and the development of full-scale biomethanation systems is expected to require a significant effort due to the configurational challenges that need to be addressed to obtain high quality of the upgraded biogas. One option to make use of the existing volume of anaerobic digesters for methanation is to develop extremely low-cost in situ biomethanation and accept only partial in situ biogas upgrading by putting it in series with an ex situ methanation reactor for an overall improvement of process costs (Jensen et al., 2020a,b).

12.5.2 Ex situ biomethanation

In the ex situ configuration (Fig. 12.4), the methanation of biogas takes place in a separate reactor. The CO2 is therefore not produced internally in the ex situ reactor, but supplied exogenously along with H2. The CO2 can be supplied both as purified CO2 and as part of raw biogas (Fig. 12.4). Although ex situ biogas upgrading requires the implementation and cost of a dedicated methanation reactor, it has received significantly more interest for biogas upgrading than in situ biomethanation. The strong interest in ex situ biomethanation is caused by the ability to design the process specifically for methanation without being compromised by the operation and design of the anaerobic digester or compromising processes of the anaerobic digestion. Since the reactor functions independently of the anaerobic digester, lab-scale experiments have tested a range of different reactor designs, with continuous stirred tank reactors (CSTRs) (Kougias et al., 2017; Seifert et al., 2014), and trickle-bed reactors being subject to the most intense development (Burkhardt and Busch, 2013; Savvas et al., 2017). These reactor types are presented in more detail in Section 12.7. The primary focus point for the ex situ reactors has been the optimization of the gas to liquid mass transfer of H2 to the residing methanogens. In contrast to in situ methanation where the natural microbiome of the biogas reactor is utilized to catalyze the methanation, both complex microbial communities and pure cultures of methanogens can be used as microbial bases in ex situ reactors. Reactors with axenic pure cultures, using only a single strain of a hydrogenotrophic methanogen, like Methanothermobacter marburgensis or Methanothermobacter thermautotrophicus, have subsequently been demonstrated (Martin et al., 2013; Seifert et al., 2014). However, most studies have been carried out with the use of complex cultures, due to the aforementioned ease of operation, low-cost inoculum, and reduced need for costly process hygiene to maintain axenic culture purity (Table 12.3). In these reactors, the exogenous addition of H2 leads to a pronounced increase in hydrogenotrophic methanogens. In thermophilic reactors based on mixed-culture inocula, species of Methanothermobacter like those used for pure culture reactors are enriched, but in contrast to the pure culture reactors, these reactors often include other methanogenic strains as well (Guneratnam et al., 2016; Kougias et al., 2017; Porté et al., 2019). Although the continued addition of H2 clearly favors the proliferation of hydrogenotrophic methanogens, ex situ reactors based on mixed-

culture inocula also entail the continued presence and activity of microorganisms involved in the fermentative processes found in the traditional anaerobic digester (Bassani et al., 2017; Jensen et al., 2019; Kougias et al., 2017). Despite its promises of low cost and ease of operation, the use of a complex inoculum results in establishment of a complex catalytic biomass that can be difficult to control, and will often catalyze unwanted reactions like the production of acetic acid from H2 and CO2, through the process of homoacetogenesis (Liu et al., 2016; Section 12.6.5).

Table 12.3

In situ/ ex situ

Methane production rate (NL/L/d)

CCH4 (%)

YCH4/H2 (%)

Reactor configuration

In situ

0.74

96

>91

Continuous stirred tank rea

In situ

0.45

78

N.A.

CSTR

In situ

0.38

65

92

CSTR

In situ

0.30

85

95

CSTR

Ex situ

690.8

23

N.A.

CSTR

Ex situ

233.6

96

96

CSTR

Ex situ

171.4a

95

>95

CSTR

Ex situ

144.0

23

96

Trickle bed

Ex situ

137.6

85

N.A.

CSTR

Ex situ

135.8

25

>97

CSTR

Ex situ

122.7

58

100

Trickle bed

Ex situ

72.0a

95

>95

CSTR

Ex situ

61.0

15

100

CSTR

Ex situ

53.8

7

N.A.

CSTR

Ex situ

41.1

14

N.A.

Trickle bed

Ex situ

33.7

44

100

Trickle bed

Ex situ

26.4

90

98

Trickle bed

Ex situ

15.4

99

99

Trickle bed

Ex situ

11.0

97

99

Trickle bed

Ex situ

9.8

61

N.A.

CSTR

Ex situ

4.0

91

92

CSTR

Ex situ

3.0

98

100

CSTR

Ex situ

2.3

83

76

Trickle bed

Ex situ

1.6

97

N.A.

Trickle bed

Ex situ

1.5

95

99

Trickle bed

Ex situ

1.5

95

99

Trickle bed

Ex situ

1.4

87

96

Trickle bed

Ex situ

1.2

99

93

Fixed bed

Ex situ

0.7

96

92

Fixed bed

Ex situ

0.4

79

N.A.

CSTR

aEstimated from available data.

In general, ex situ biomethanation expands the possibilities to control the reactions compared to the in situ configuration. However, most of the factors controlling the efficiency of the methanation reaction are the same for both process types. We will discuss the dominant controlling factors in the next section.

12.6 Factors controlling biomethanation

Like microbial processes in all other natural and technical biological systems, a range of factors control the rate of reaction and CH4 yield in both in situ and ex situ biomethanation reactors. The main controlling and influencing factor in biomethanation is H2 gas–liquid mass transfer, which governs the supply of H2 to the methanogens and thereby the methanation reaction rate. However, microbial activity plays an essential role in mass transfer and certain biological conditions, including nutrients, temperature, pH, and CO2, and must therefore be satisfied to sustain high methanogenic activity and secure a stable process with limited activity of competing microorganisms. These factors are discussed in the subsequent sections.

12.6.1 Mass transfer of H2

Since methanogens require water for maintaining cellular activity, they are inherently present in aqueous or moist environments and can therefore only convert dissolved H2. The low aqueous solubility of H2 makes gas–liquid mass transfer a key factor in regulating the activity of methanogens and competing hydrogenotrophic microorganisms that may be present. The development of new technologies for biomethanation has therefore for a large part focused on optimizing the rate of H2 gas–liquid mass transfer (i.e., the rate of H2 solubilization), which has been identified as the main controlling factor of the reaction rate in biomethanation (Lecker et al., 2017). Gas solubility in dilute aqueous solutions can be described by Henry’s law [Eq. (12.5)], which states that the amount of gas that can be dissolved in liquid is directly proportional to its partial pressure in the bulk gas phase above the liquid (Cussler, 1997).

(12.5)

where Cl* is the equilibrium concentration of the gas dissolved in the liquid phase, H is the Henry’s law constant (or the proportionality factor), which can either be calculated or found empirically, and Pg is the gas partial pressure above the liquid. Henry’s law constants for some of the most relevant gases in biomethanation processes are listed in Table 12.4. Apart from the reactants and products of methanation (H2, CO2, CH4), the table also includes oxygen (O2) which is an inhibitor of the methanogens, and ammonia (NH3) which is an essential part of the buffer system in the anaerobic digester system.

Table 12.4

Gas

H (bar/mol/kg)

Ionic form(s) in water

Hydrogen (H2)

1282

No

Carbon dioxide (CO2)

29

Yes

Methane (CH4)

769

No

Oxygen (O2)

769

No

Ammonia (NH3)

24.4×10−5

Yes

Providing enough substrate to the methanogens (i.e., H2 and CO2) is essential to maintain a high CH4 production rate. From Table 12.4 it can be inferred that gaseous CO2 is approximately 44 times more soluble than H2. However, the acid–base properties of CO2 mean that its overall dissolution capacity is pHdependent and therefore can be several fold higher than predicted by Henry’s law due to its potential transformation to carbonic acid, bicarbonate, and carbonate in aqueous solutions [Eq. (12.6)].

(12.6)

The total pool of dissolved CO2 including its ionic and protonated forms is termed dissolved inorganic carbon (DIC). By including the total DIC pool, the total CO2 absorption capacity is thus 1.5–5 times higher within the common pH range of biomethanation reactors (pH 6–8), than that directly inferred from Henry’s law. CO2 limitations are therefore generally not thought to be an issue in biomethanation reactors (Chen et al., 2019). The low solubility of H2 is conversely a clear barrier for supplying H2 to the methanogens—a challenge which is accentuated by the fact that methanation requires four moles of H2 per mole of CO2 for the formation of one mole CH4 [Eq. (12.1)]. Since the low aqueous solubility of H2 is directly constraining the availability of H2, rapid mass transfer of H2 from the gaseous to the liquid phase is a critical parameter of biomethanation. H2 gas–liquid mass transfer is a physical phenomenon that takes place when the gas phase and dissolved H2 concentrations are not in equilibrium, which in an active biological system is caused by the microbial consumption of H2. The forces that govern the mass transfer of H2 are universal in all biomethanation reactor configurations and designs: Whether H2 is supplied as bubbles created by sparging in a CSTR (Bassani et al., 2017) or supplied by diffusion through a gas–liquid interface to a biofilm in a trickling-bed reactor (Burkhardt and Busch, 2013) (Fig. 12.5). The following paragraphs describe how these physical factors affect gas–liquid mass transfer and put them in the context of selected biomethanation process conditions to show how H2 gas– liquid mass transfer and therefore the biomethanation reaction rate can be influenced by reactor design and operation.

Figure 12.5 Mass transfer of H2 from gas to liquid phase in common biomethanation reactor types: (A) Through the gas–liquid interface of a H2 gas bubbles in a continuous stirred tank reactor. (B) In a biofilm reactor across a stagnant liquid layer to an active biofilm of methanogens.

The transport of H2 from the bulk gas phase (e.g., a bubble) to the bulk liquid phase can be described by three physical mechanisms, advection, turbulence, and diffusion (Fenchel et al., 1988) (Fig. 12.6):

Advection is the directional and orderly transport of molecules due to bulk fluid motion. Turbulence is random motion of eddies and small vortices, which provides local mixing of bulk gas and liquid. Diffusion is microscopic transfer of molecules from high to low concentration regions until equilibrium is reached. Diffusion takes place in stagnant gas and liquid films adjacent to the gas–liquid interphase, and constitutes the ratelimiting process in biomethanation.

Figure 12.6 The two-film theory, describing molecular transport over the gas–liquid interface. In the bulk gas and liquid phases, advection and turbulence ensure mixing and bulk phases are assumed to be homogeneously mixed with no gradients. Concentration gradients drive the diffusion in the stagnant film layers. Concentrations in the gas phase are described by: P, the gas partial pressure in the bulk gas phase; Pi, the gas partial pressure at the gas side of the gas–liquid interphase; kg, the gas-side mass transfer coefficient for gas diffusion through the stagnant gas–phase layer. Concentrations in the liquid phase is described by: Ci, the concentration of solubilized gas at the gas–liquid interphase; C the concentration of solubilized gas in the bulk liquid phase; kl, the liquid-side mass transfer coefficient for diffusion of the solubilized gas through the stagnant liquid zone.

The overall rate of H2 gas–liquid transfer across a gas–liquid interface can be described by the two-film theory (Lewis and Whitman, 1924). The theory formulates a simplified mathematical expression to describe molecular mass transfer from bulk gas to bulk liquid. The two-film theory divides the overall transport from gas to liquid into five zones (Fig. 12.6B):

1. Bulk gas, where the gas is assumed to be completely mixed. 2. A stagnant gas film, where gas–gas diffusion is the dominant transport process. 3. Gas–liquid interphase, where the gas and liquid zones meet. The interphase has no volume and gas and liquid phases are assumed to be in equilibrium according to Henry’s law. 4. A stagnant liquid film, where gas–liquid diffusion is the dominant transport process. 5. Bulk liquid, assumed to be completely mixed.

H2 diffusion through the liquid stagnant layer is broadly acknowledged as the rate-limiting step for H2 gas–liquid mass transfer in biomethanation systems. The rate of diffusion through a gas–liquid interface (i.e., diffusional flux) in an aqueous media can be described by Fick’s law (Fenchel et al., 1988). The equation here expresses the flux of H2 through water [Eq. (12.7)].

(12.7)

where is the absolute molar flux of H2 (mol/s/m²), is the diffusion coefficient of H2 in H2O (m²/s), is the concentration gradient of H2 (mol/m³/m), where x denotes the depth of the H2 penetration into the stagnant diffusive layer. The magnitude of the diffusion coefficient is determined by the properties of the solute (in this case H2) and the solvent. The negative sign denotes the flux from higher to lower concentration, a gradient which in biomethanation systems is created and sustained by methanogenic conversion of the transferred H2. The net movement of H2 from the gaseous phase to the methanogens in the aqueous phase in biomethanation reactors is therefore fueled by the consumption of dissolved H2, similar to how microbial activity influences the driving force for molecular diffusion in both natural ecosystems, technical systems like wastewater treatment plants, and biological air filters (Fenchel et al., 1988). The two-film theory can be related to Fick’s first law [Eq. (12.7)] to describe the diffusional flux through the individual stagnant gas [Eq. (12.8)] and liquid [Eq. (12.9)] films as the product of the diffusional gradient (i.e., the concentration difference between a bulk phase and the gas–liquid interface) and a film-specific mass transfer coefficient. Gas film flux

(12.8)

Liquid film flux

(12.9)

J is the molar flux of gas (mol/m²/s), and kg and kl are the local mass transfer coefficients (or rate constants, m/s) for the transport across the gas and liquid layers, respectively. P is the partial pressure in the bulk gas phase, C is the concentration of dissolved gas in bulk liquid. The Pi and Ci are the concentrations of gas in the gas and liquid phase at the gas–liquid interphase, respectively, and these concentrations are assumed to be in equilibrium. kg and kl, as defined by the two-film theory, are directly derived by integrating Fick’s first law with a constant concentration gradient through the stagnant film:

(12.10)

where k denotes kg or kl (m/s), is the diffusion coefficient of H2 in gas ( ) or liquid ( ), and x is the thickness of the stagnant film (gas or liquid). Eq. (12.10) implies that the molar flux through a stagnant film layer depends on the diffusion coefficient of the solute and the thickness of the diffusional boundary layer. Interfacial concentrations (Pi and Ci) are not readily determined, but an overall mass transfer coefficient linking flux to measurable concentrations of gas in the bulk gas and liquid phases can be defined similarly to the expressions in Eqs. (12.8) and (12.9), by assuming that the flux is in steady state. The overall mass transfer coefficient (Kl) is a function of the film-specific mass transfer coefficients [Eq. (12.11)].

(12.11)

Eq. (12.11) shows that the main resistance to H2 gas–liquid mass transfer lies in the liquid film, because of low H2 solubility (i.e., high Henry’s constant, Table 12.3), which makes the gas-side diffusional resistance negligible (Table 12.4). The overall H2 gas–liquid mass transfer rate in a given reactor is furthermore linked to the flux by the available gas–liquid interfacial area over which the solute can be transported, and the overall expression for the H2 gas–liquid mass transfer rate can therefore be expressed as:

(12.12)

where is the H2 gas–liquid mass transfer rate (mol/m³/s), is the liquid side mass transfer coefficient used as an approximation for the overall mass transfer coefficient ( ) (m/s), is the volumetric gas–liquid interfacial area (m²/m³), is the concentration of dissolved H2 in equilibrium with its gas phase partial pressure [Eq. (12.5)] (mol/m³), and is the actual concentration of dissolved H2 in the liquid phase. Several operational factors affect the individual in Eq. (12.12), which accordingly has formed the basis for development of reactor design and operation for biomethanation, as exemplified in the following.

12.6.1.1 H2 partial pressure in gas phase

Increasing the partial pressure of H2 will increase the equilibrium concentration of dissolved H2 [Eq. (12.5)] and thus the concentration driving force for H2 gas– liquid mass transfer [Eq. (12.12)]. The increase in pressure is, for example, utilized by the company Electrochaea, which has successfully demonstrated the application of biomethanation technology for upgrading of biogas at demo-scale by operating at a headspace pressure of 10 bar(g) (Electrochaea, 2018).

12.6.1.2 Interfacial area

The gas–liquid interfacial area is critical for gas–liquid mass transfer, since transfer rates are proportional to the available surface [Eq. (12.12)]. The gas– liquid interfacial area can be managed and influenced in different ways depending on the reactor type. The way of generating a gas–liquid interfacial area is a primary factor underlying the selection of reactor type (Section 12.7). For trickle-bed reactors, the biofilm-colonized surface area of the carriers can be optimized to as high as possible, similar to biofilters employed for removal of odorous compounds (Andreasen et al., 2013). In CSTRs, the H2 is often introduced as gas bubbles. The gas–liquid interphase can here be optimized by optimizing bubble formation and dispersion (Bassani et al., 2017), and through continuous stirring to break up bubbles and keeping them in suspension (Peillex et al., 1990).

12.6.1.3 Methanogenic activity

For practical purposes, the actual concentration of H2 in the bulk liquid phase can often be assumed to be close to zero due to the low solubility of H2 and the fast biological consumption within a narrow liquid layer of 0.5–1 mm adjacent to the gas–liquid interphase (Garcia-Robledo et al., 2016; Maegaard et al., 2019). Eq. (12.10) shows how the mass transfer coefficient is influenced by the thickness of the diffusional stagnant layer, as modeled by the two-film theory.

However, the two-film theory does not take H2 conversion in the stagnant liquid layer into . Actual methanogenic activity in the boundary layer would effectively decrease the distance a H2 molecule has to diffuse to reach the layer of microbial activity compared to what is described by the two-film theory. Increasing the biological potential for H2 consumption is therefore a continued point of optimization. The potential for biological optimization can be witnessed by the fact that the identified methanogenic community in some biomethanation studies only constitutes 4%–8% of the total microbial community (Bassani et al., 2017; Kougias et al., 2017). The company Electrochaea has optimized the biological effect by using a pure methanogenic culture to increase the concentration of the active catalytic biomasses. The cells will therefore not only be present in bulk liquid, but also in the stagnant liquid layer, which will thereby decrease the liquid side resistance dramatically (see Fig. 12.6). Enhancing the microbial activity is therefore of utmost importance for the efficiency of the biomethanation technology. The optimization of the biological part of biomethanation will be covered in the following sections.

12.6.2 Temperature

Temperature affects gas–liquid mass transfer through its effect on H2 diffusion, solubility, and biological activity, and it is therefore an important parameter in biological methanation. Increasing temperature will decrease the aqueous solubility of H2, but will also increase H2 diffusivity and thus the mass transfer coefficient [Eq. (12.10)]. It has been hypothesized that the higher diffusion speed of H2 in water can outweigh the decrease in solubility, resulting in an overall higher mass transfer of H2 with increasing temperatures (Jensen et al., 2020b). As for all chemical and biological reactions, methanogenic reaction rates increase with increasing temperature. The higher mass transfer rate of H2 is therefore also followed by increased biological conversion rates, which may further enhance the gas–liquid mass transfer of H2. Methanogens are generally grouped into species that can grow within the mesophilic growth range of 37°C– 45°C, or the thermophilic growth range of 55°C–70°C, which reflects the operating conditions chosen for the biomethanation systems that are either mesophilic or thermophilic (Zabranska and Pokorna, 2018).

Thermophilic biomethanation systems generally exhibit higher CH4 production rates than mesophilic systems (Jensen et al., 2020b). In a study comparing methanation at different temperatures, Luo and Angelidaki (2012) reported a doubling in CH4 production rates in an ex situ thermophilic (55°C) CSTR compared to operation at mesophilic (37°C) conditions (Luo and Angelidaki, 2012). A higher methanation reaction rate reduces the reactor volume needed for converting a given amount of CO2 and H2 to CH4. From an economic point of view, the advantage of reduced capital expenses (CAPEX) for a smaller bioreactor has to be balanced against the increase in operational expenses (OPEX) for the energy needed for operation at higher temperatures. At least a part of this heating requirement can potentially be covered by the reaction heat generated from the exothermic methanation reaction (Jensen et al., 2020a).

12.6.3 Growth requirements

For the methanogens, their conversion of H2 and CO2 to CH4 has the sole purpose of harvesting energy for cellular processes like cell maintenance and cell division, where some of the CO2 is assimilated into cellular biomass. A study on a pure culture of the methanogen Methanobacterium congolense showed that 11%–26% of the consumed CO2 is assimilated as biomass, giving an approximate one mole of CO2 incorporated into biomass for every 4 moles of CH4 produced (Chen et al., 2019). However, the exact coupling between CH4 formation and growth has been found to depend on both the organisms, as well as the growth conditions like the concentration of dissolved H2 (de Poorter et al., 2007, and references herein). Dupnock and Deshusses (2017) proposed the following reaction scheme to combine the methanation reaction [Eq. (12.1)] with methanogenic growth [Eq. (12.13)].

(12.13)

The overall reaction stoichiometry as proposed here suggests that the maximum yield of CH4 from CO2 and H2 through biological methanation is 85% and 92% of the theoretical yield as proposed by Eq. (12.1), respectively, which align well with general CH4 yields observed in biomethanation studies (Jensen et al., 2020b). H2 assimilation into biomass growth represents a loss of energy, but can be seen as a necessary energy input needed to maintain the reaction catalysts. Part of the “lost” energy used for biomass growth can furthermore be regained through the process of anaerobic digestion following the processes outlined in Section 12.4.2. This can occur either directly in the methanation reactor if it is based on a mixed culture (in situ or ex situ) and therefore inevitably contains a diversity of heterotrophic bacteria that can degrade organic matter (i.e., biomass); or by introducing biomass excreted from the methanation reactor to the conventional anaerobic digester from where the methanation reactor already receives its CO2 substrate gas. The reaction scheme in Eq. (12.13) is simplified in the sense that methanogenic activity and proliferation are conditional on the presence of several additional elements. These elements are inorganic nutrients, organic growth factors, and vitamins essential for methanogenic growth and activity. Macronutrients like nitrogen, phosphorous, and sulfur form integral parts of amino acids, nucleic acids, etc., while micronutrients like cobalt, iron, and copper function as cofactors in enzymes (Madigan et al., 2019b). The main elements of methanogenic cells are depicted in Table 12.5.

Table 12.5

Element

Methanogenic cells

Carbon

37%–44% (w/w)

Oxygen

N.D.

Nitrogen

9.5%–12.8%

Hydrogen

5.5%–6.5%

Phosphorous

0.5%–2.8%

Sulfur

0.56%–1.2%

Potassium

0.13%–5%

Sodium

0.3%–4%

Calcium (Ca)

85–4500 ppm

Magnesium (Mg)

0.09%–0.53%

Iron (Fe)

0.07%–0.28%

Nickel (Ni)

65–180 ppm

Cobalt (Co)

10–20 ppm

Molybdate (Mo)

10–70 ppm

Zinc (Zn)

50–630 ppm

Copper (Cu)

<10–160 ppm

Manganese (Mn)

<5–25 ppm

Although the nutrients are required in different amounts, they are all essential to methanogenic cell function and can therefore all limit methanogenic activity if not present in the liquid medium in sufficient concentrations. For the in situ biomethanation which is carried out in the anaerobic digester, the medium is always the digestate of the reactor. It is assumed that all the necessary nutrients are present in the digestate, because it is continuously being partly renewed by fresh biomass that is typically obtained from a variety of different sources. The methanation reaction rate in in situ biomethanation is consequently assumed never to be limited by nutrient shortages. This is in contrast to ex situ reactors, where the liquid medium is supplied exogenously to the reactor; the medium’s composition can therefore be determined by the operator. Some ex situ systems use digestate-derived solutions because they constitute a simple and low-cost nutrient source. Utilized nutrient solutions include degassed digestate (Corbellini et al., 2018), centrifuged and filtered digestate (Savvas et al., 2017), pasteurized cattle manure (Sieborg et al., 2020), or reject water from dewatered sludge of an anaerobic sludge digester (Jensen et al., 2019). The use of digestate-derived nutrient medium also means that (1) its chemical composition is largely uncontrolled; (2) organic material is continuously introduced to the reactor, which inherently leads to proliferation of bacteria involved in organic matter degradation; and (3) competing microorganisms with potential adverse effects on reactor stability and CH4 yield are continuously introduced to the methanation reactor (see Section 12.6.5). Many ex situ studies therefore use a synthetic medium where nutrients have been added as mineral salts (NH4Cl, KH2PO4, MgCl2, CaCl2, etc.) in known concentrations (Dupnock and Deshusses, 2019; Jee et al., 1987; Rachbauer et al., 2016). The design and mixing of synthetic media is more costly in relation to both labor and chemicals, but allows for a higher degree of control of the process. There is currently no consensus on which medium constitutes the overall best nutritional source in ex situ biomethanation reactors: Digestate-based solutions stand as the immediately cheapest nutrient source, but their usage introduces more process complexity because the composition is often not known in detail. This added complexity can be circumvented by using synthetic media, where the exact composition is known.

12.6.4 pH and CO2

Many of the processes in methanation reactors influence the pH of the reactor by either producing or consuming acids and bases, which makes pH a crucial point of process control. Methanogenesis takes place in the pH range of 6.5–8.5, with an optimum at pH 7–8 (Weiland, 2010). This relatively narrow range makes methanogens sensitive to both high and low pH, underlining pH control as a key element in biomethanation reactors. Many of the chemical elements in the nutrient solutions—digestate-based or synthetic—function not only as nutrients for the microorganisms but also impact pH and buffer capacity. The dominant acid–base pairs controlling pH in digestate from animal excreta are CO2/HCO3−/CO3²− [Eq. (12.6)], NH4+/NH3, and CH3COOH/CH3COO−, and only to a minor degree Mg²+, Ca²+, and inorganic phosphates (Sommer and Husted, 1995). In addition to these acid–base pairs, organic macromolecules in the digestate will have buffering capacity as well. In synthetic media, the dominating buffer systems can be specifically tailored and often consist of phosphate salts, but also here, the pH is influenced by the CO2 reactant and NH3 supplied with the biogas. In biomethanation systems, the pH and buffer effect is furthermore influenced by the production of metabolic water that inherently accompanies CH4 production [Eq. (12.1)] and results in dilution of the buffer systems. This effect is especially prominent in trickle-bed reactors, where the volume of supplied medium is much lower compared to submerged reactor systems and the metabolic water consequently contributes more to the total liquid content as a consequence (Strübing et al., 2017). The buffer of primary importance for controlling pH in biomethanation systems is the pool of DIC represented by CO2/HCO3−/CO3²−, since this is the most prone to change due to CO2 acting as both substrate and buffer. If CO2 conversion rates become higher than the CO2 feed or production rate in a biomethanation reactor, the DIC will start to deplete according to Le Chatelier’s principle [Eq. (12.6)]. The main fraction of dissolved CO2 will be in the form of HCO3− during common pH conditions of biomethanation reactors (pH 7–8), so CO2 depletion is accompanied by H+ depletion, which causes pH to increase. CO2 depletion can ultimately result in pH increases to levels unsuitable for methanogenic activity with an adverse impact on the overall process performance (Agneessens et al., 2017; Luo et al., 2012; Wahid et al., 2019). The

exact pH value at which CH4 production rate start to decrease has been found to be different for each of these studies: pH 8.18 (Agneessens et al., 2017), pH 8.3 (Luo et al., 2012), and pH 8.5 (Wahid et al., 2019). The latter study reported an increase in pH up to a maximum of pH 9.4 (Wahid et al., 2019). Inhibition of methanogenesis at pH 8.0–8.5 corresponds with the general optimum for methanogens of pH 7–8 (Weiland, 2010). The dual role of CO2 as both substrate and buffer is complicated by the fact that the end product of methanation requires very low concentrations of CO2 to comply with the specifications of natural gas (Table 12.1). The ratio of H2 and CO2 in the feed gas thus needs careful adjustment to the actual reaction stoichiometry in the reactor to satisfy the requirements of product gas quality without depleting dissolved CO2. The actual reaction stoichiometry inherently deviates from the theoretical reaction [Eq. (12.1)] due to microbial growth and maintenance [Eq. (12.13)], and may vary over time due to changes in methanogenic growth conditions, or the activity of competing microorganisms (see Section 12.6.5). Careful CO2 balancing can be circumvented by applying more active pH regulation with acids (Luo and Angelidaki, 2013), or the addition of synthetic medium containing other buffer systems (Strübing et al., 2017). In the lower pH range of methanogenic activity, a decrease in gas production to 85% of its basis CH4 production rate was observed when pH in the medium temporarily decreased to pH <6.2 (Strübing et al., 2017). The decrease in pH observed by Strübing et al. (2017) could potentially be attributed to the formation of VFAs, predominantly the formation of acetate via the process of homoacetogenesis. Regardless of the direction of the pH imbalance, the effect of too acidic or too basic pH is reversible, although the original methanogenic activity may not be regained instantaneously (Maegaard et al., 2019). However, the degree of reversibility must be expected to depend on the time duration of unfavorable pH conditions and could in extreme cases require regrowth of the methanogenic community for the full resumption of activity.

12.6.5 Bacterial interaction and competition

The use of pure methanogenic cultures and synthetic media allows for a high

degree of process control, not least because process disturbance by the introduction of competing microorganisms with inoculum and medium can be avoided. In in situ methanation, the use of complex cultures is inherent, because of the continuous load of organic material to the anaerobic digester. However, even for ex situ configurations, the use of complex cultures is for many developers the preferred choice due to issues of cost and robustness related to the conditions needed to maintain a pure methanogenic culture (Table 12.3). As described in Section 12.4.2, the anaerobic digestion of organic material includes a range of different interdependent microbial processes. One of the reported effects is competition for substrate (H2, CO2) by homoacetogenic bacteria as their conversion of H2 and CO2 into acetate [Eq. (12.4)] constitutes direct competition to the desired activity of hydrogenotrophic methanogens that use the same substrates for production of CH4 [Eq. (12.1)]. The homoacetogenic bacteria are inherently present in mixed culture-based biomethanation reactors where they utilize a range of substrates, however, when H2 is introduced in high concentrations, these acetogens can dominate H2 consumption. Experiments by Liu et al. (2016), showed that increased influx of exogenously added H2 caused the activity of homoacetogens to increase (Liu et al., 2016). They also reported that up to 40% of exogenous H2 added to an anaerobic digester was consumed by homoacetogens compared to 2%–5% of the H2 produced from the degradation of organic matter under standard operating procedures. The reported dominance of homoacetogens following the addition of H2 is ed by Jensen et al. (2019) who found that up to 63±8% of the added H2 was converted to acetate in a biofilm-based reactor setup. Homoacetogenic acetate production is indicated through elevated acetate concentrations in many biomethanation studies and seems to be common to both in situ and ex situ configurations based on mixed-culture inocula (Agneessens et al., 2018; Jensen et al., 2019; Rachbauer et al., 2016). However, the dominance of homoacetogenesis and the associated accumulation of acetate has been shown to be transient and most pronounced during start-up of methanation reactors and potentially during changing operational parameters, for example, organic loading rate during in situ methanation (Agneessens et al., 2018; Rachbauer et al., 2016). The dominance of methanogens over homoacetogens after continued addition of H2 could potentially be explained by the energy yield of hydrogenotrophic methanogenesis and acetogenesis, respectively. From a thermodynamic point of view, the methanogens have a growth advantage, reflected in the lower Gibbs energy under standard conditions for hydrogenotrophic methananogenesis (ΔG° ′=−131.0 kJ/mol) compared to homoacetogenesis (ΔG°′=−94.9 kJ/mol) (Schink,

1997). On an enzyme level, the substrate affinities also indicate that methanogens are able to efficiently utilize lower concentrations of dissolved H2 compared to homoacetogens. For the homoacetogenic strain Acetobacterium woodii, the threshold for H2 utilization with CO2 as a terminal electron acceptor was found to be 250 Pa (Poehlein et al., 2012). In comparison, the H2 concentration threshold for methanogenic species from different genera was: Methanobrevibacter arboriphilus ~8 Pa, Methanosarcina barkeri 150 Pa, M. marburgensis ~8 Pa (Kaster et al., 2011). The methanogens are therefore suited to exploit lower levels of dissolved H2, which is an advantageous trait when considering the low solubility of H2. Although several studies report accumulation of acetate, no studies report reactor failure as a direct consequence of the increased acetate levels. However, the use of acetic acid as an antimicrobial agent for preservation of food is well known, and the levels at which acetate could limit the methanogenic activity is therefore of interest to determine the robustness of the technology. To elucidate at which acetate levels the methanogenic activity is inhibited, Rachbauer et al. (2017) performed batch assays using biomass from a trickle-bed reactor amended with 400–1200 mg/L at pH 5.7. Even at 400 mg/L the acetate was found to be inhibitory and resulted in a further decrease to pH 5.3–4.2 in the incubations (Rachbauer et al., 2017). In contrast, Kougias et al. (2017) observed accumulation of 1600–4000 mg/L acetate and reactor pH of 8.3–8.5, with only a limited effect (Kougias et al., 2017). A probable explanation for this difference can be deduced from knowledge on the mechanisms of toxicity of organic acids like acetic acid, where it is only the protonated form of acetate (acetic acid) which can cross the cellular membrane and disrupt processes internally in the microbial cell (Trček et al., 2015). The correlation between acetate level and pH is therefore essential when considering inhibitory levels of organic acids. The equilibrium between the protonated and unprotonated can be calculated using the Henderson–Hasselbalch equation [Eq. (12.14)].

(12.14)

where pKa is the pH value at which the protonated acid (HA; acetic acid) and

unprotonated acid (A−; acetate) are present at a 1:1 ratio. The pKa for acetic acid is 4.75. Considering this and maximum acetate concentrations of 4000 mg/L (pH 8.3) obtained by Kougias et al. (2017) to the lowest inhibitory concentration tested of 400 mg/L (pH 5.7) obtained by Rachbauer et al. (2017), the concentration of protonated acetic acid is 40 times higher in the latter, even though the total concentration of acetate is 10 times lower. In the same way, it can be calculated that for the highest level of 1200 mg/L tested by Rachbauer et al., the concentration of acetic acid is 119 times higher. The levels of acetate and its potential detrimental effect on the reactor microbiology should therefore always be evaluated in the context of pH as well. Homoacetogenesis affects both the production of acetic acid and reduces the media pH, which can reduce the process efficiency if left uncontrolled. To negate the effects of acetic acid accumulation, control of reactor pH is of critical importance. pH can be controlled through liquid media buffering, control of CO2 conversion, dilution, and base dosing. pH control is integrated in most reactor setups and a limited effect of acetic acid accumulation is therefore observed during normal operation.

12.7 Reactor design for biological methanation

The optimal design and operation of a biomethanation reactor facilitates high H2 gas–liquid mass transfer rates to the active methanogenic community. Independent of whether the methanogens are suspended in liquid or fixed in a biofilm, the process conditions should facilitate the establishment of only a thin liquid diffusional layer, and a high concentration of methanogens that can convert dissolved H2 instantly to maximize diffusional rates. The CSTRs and trickle-bed reactors are covered in this section, as these are the best studied reactor types and at the highest technological readiness level. Only studies where biomethanation reactors have been operated in continuous mode are summarized in Table 12.3.

12.7.1 Continuous stirred tank reactor

The CSTR is the most frequently used reactor technology for bioprocesses and is also the preferred reactor type for biogas production and biomethanation. CSTRs have been used for both in situ and ex situ biomethanation, and are designed with a stirrer system and a gas diff which together provide and disperse H2 bubbles in the liquid broth. The diff is designed to supply H2 as small bubbles, characterized by a large surface–volume ratio, to enhance the gas– liquid interface and thus diffusion of H2 into the liquid. Stirring keeps both biomass and bubbles in suspension in the reactor and plays a vital role in the H2 gas–liquid mass transfer by (1) mechanically breaking up gas bubbles to increase the gas–liquid interfacial area and (2) creating mixing patterns that prolong the gas retention time in the liquid phase (Garcia-Ochoa and Gomez, 2009). CH4 production has consequently been shown to increase with stirring intensity due to increased H2 gas–liquid mass transfer. Increasing mixing intensity from 150 to 300 rpm was, for example, shown to increase the H2 conversion rates and concomitantly the CH4 content in the outlet from 53±3% to 68±2.5% in a lab-

scale in situ study (Luo and Angelidaki, 2013). In the same study, it was shown that decreasing pore size of the diff, from a ceramic diff with pore size of 0.5–1.0 mm to a column diff with a pore size of 14–40 µm resulted in increased CH4 content in the product gas from 68% to 75% due to dispersion of smaller H2 bubbles. The CSTR concept is utilized by two companies within the biomethanation industry, Electrochaea GmbH and MicrobEnergy GmbH. Both systems supply H2 by sparging and active stirring, which along with elevated pressure and temperature facilitate optimal transfer and conversion of H2. Electrochaea has established the first biomethanation pilot plant at Avedøre Wastewater Treatment Plant near Copenhagen, Denmark. The CSTR is operated at a reactor volume of ~7 m³ of active volume (Electrochaea GmbH, 2017). To facilitate mass transfer, the Electrochaea technology is operated at 10 bar(g) and 63°C (Electrochaea, 2018). The MicrobEnergy CSTR plant at Allendorf, , has a size of 5 m³, and is also operated at elevated pressure and temperature: 5–15 bar and 50°C–80°C, respectively (IEA, 2018). The challenge of gas–liquid mass transfer in CSTR systems is not unique to the biomethanation technology but known from fermentation processes where O2 gas is supplied as electron acceptor. O2 is a sparingly soluble gas similar to H2 (Table 12.4), and high O2 gas–liquid mass transfer rates are critical to product formation in fermentation systems, in the same way that H2 gas–liquid mass transfer is important to product formation in biomethanation. A lot of the process knowledge that has been obtained within O2 gas–liquid mass transfer can thus be adapted for biomethanation purposes. Due to the industrial importance of fermenters, a range of correlations exists for the gas–liquid mass transfer of O2 in biological fermenters (reviewed by Garcia-Ochoa and Gomez, 2009). These correlations are often derived from empirical data. One of the most widely used correlations to estimate kLa during O2 gas–liquid mass transfer in CSTRs was formulated by Van’t Riet (1979) in a system based on pure water (Garcia-Ochoa and Gomez, 2009; Van’t Riet, 1979):

(12.15)

where Vs is the superficial gas velocity; P/V is the power–volume ratio. The

exponents K, a, and b are empirically derived from reactor experiments. Empirical mass transfer correlations are dependent on a range of parameters, including reactor geometry, scale, and liquid characteristics such as viscosity, density, and surface tension. The Van’t Riet correlation is therefore only valid in water but nonetheless exemplifies the influence of some of the main operational parameters on gas–liquid mass transfer in CSTRs: the gas velocity (Vs), which depends on reactor diameter and gas flow rate; and the volumetric power input, which directly relates stirring intensity to reactor scale. The Van’t Riet correlation [Eq. (12.15)] hence shows that gas–liquid mass transfer in a CSTR biomethanation reactor can be promoted either by increasing the H2 gas flow rate or by increasing the stirrer intensity. However, with the specifications to the allowed H2 concentration in the product gas in mind (Table 12.1), there will be a limit to how much the H2 gas flow rate can be increased. This is the main difference between O2 gas–liquid mass transfer in fermentations, and H2 gas– liquid mass transfer in biomethanation. In industrial fermenters, a high level of dissolved O2 needs to be maintained to secure proper microbial metabolism and thus product quality and yield (Chen and Wilde, 1991). The products from these processes are present in the liquid broth and their quality is thus not affected by unconverted O2 in the off-gas. For the methanation system, full conversion of the supplied H2 and CO2 is conversely essential, because the outlet gas is the process product, and only limited concentrations of H2 and CO2 are tolerated here (Table 12.1). The H2 gas flow rate can thus only be used to promote H2 gas–liquid mass transfer to the extent that all H2 is still being converted in the biomethanation reactor. The use of power input as a main driver for H2 gas– liquid mass transfer in CSTRs has led to the often-stated assumption that CSTRbased biomethanation reactors are associated with a significant power input for stirring (Savvas et al., 2017; Strübing et al., 2017). However, while power input inherently must constitute a larger fraction of overall operational costs in CSTRs, it is noteworthy that the only existing pilot plant biomethanation reactors are based on CSTRs. Although CSTRs constitute the most mature biomethanation reactor platform, research and development is currently addressing how to decrease the projected cost of methanation operation that the CSTR systems represent. One of the most promising systems promoting low operational cost is the trickling bed reactor presented in the next section. The technology is not at the same technological readiness level as CSTR, but is considered to be a promising low-cost technology for producing CH4 of natural gas quality.

12.7.2 Trickle-bed reactors

In the same way as the use of CSTRs for biomethanation was inspired by classical fermentation systems, the development of trickle-bed reactors stems from biological gas treatment systems, which often need to remove gas pollutants present in low concentrations from large gas volumes (Delhoménie and Heitz, 2005). A trickle-bed reactor consists of a column packed with a carrier material possessing a high volumetric surface area whereupon the microbial catalysts are immobilized. The main difference between a trickle-bed reactor and the other reactor types that have been examined for biomethanation is that the trickle-bed reactor is filled with gas instead of liquid. Having gas as the primary phase in the reactor (1) enables easy control of gas retention time without needing additional power input to achieve satisfactory product gas quality, and (2) decreases the diffusional liquid boundary layer that creates resistance to H2 gas–liquid mass transfer. An aqueous nutrient solution is occasionally trickled over the packing material to supply the basal chemical elements needed for growth and activity (Section 12.6.3). The liquid forms a thin liquid film around the biofilm, thereby creating a three-phase gas–liquid–biofilm system in the reactor. The gas phase will fill up almost the entire reactor, the liquid phase will be bound to carriers and biofilm, and the biofilm itself will be growing on the surface of the carriers in the reactor. Overall, the operation and design of trickle-bed reactors include several immediate advantages in of H2 gas–liquid mass transfer compared to liquid-based reactors: Immobilization (1) increases the retention time of the methanogenic community due to its independence from the hydraulic retention time, and (2) enables a high methanogenic surface density at the gaseous interphase. Carrier materials providing a high specific surface area are therefore expected to increase the area between the biofilm and the gas phase and thus increase the total mass transfer of H2. The trickle-bed design has been used in combination with pure cultures (Dupnock and Deshusses, 2019; Jee et al., 1987), but the inoculation and operation with pure cultures demands rigorous reactor control. Dupnock and

Deshusses (2019) inoculated a trickle-bed reactor with Methanospirillum hungatei, but after only 50 days another methanogenic species, M. arboriphilus, dominated the methanogenic community, although it comprised only 9% of the total microbial community, with no reported detection of M. hungatei (Dupnock and Deshusses, 2019). The study by Dupnock and Deshusses (2019) illustrates the stringent process conditions needed when operating reactors with pure cultures—especially in a system where nonsterile gas is introduced into the system, carrying a diverse microflora from the main anaerobic digester. For this reason, most reactors are inoculated with complex cultures, using digestate from anaerobic digesters treating either wastewater sludge or agricultural waste products (Burkhardt and Busch, 2013; Dupnock and Deshusses, 2017). The high retention time of the microorganisms makes trickle-bed reactors well suited for mixed cultures, as the hydrogenotrophic methanogens will eventually become the dominating species in the culture when H2 and CO2 are supplied as the sole energy and carbon sources. Exemplifying this enrichment, Dupnock and Deshusses (2017) used a mixed-culture inoculum and found that 27% of the retrieved sequences could be associated with Archaea after 177 days of operation with H2 and CO2, and that these were dominated by sequences affiliating with Methanobacterium and M. arboriphilus (Dupnock and Deshusses, 2017). As already mentioned, trickling bed reactors have traditionally been employed as environmental technologies for the biological treatment of waste air and gases, including organic carbon, nitrogenous, and sulfurous compounds (Delhoménie and Heitz, 2005; Fortuny et al., 2011; Ottosen et al., 2011). However, one of the main differences of these treatment technologies and the methanation reactors is the gas flow rate and composition of the inlet gas. Whereas the environmental technologies often treat large volumes of air or gas with very low concentrations of pollutants, the gas flows in the methanation reactors are often significantly lower due to the requirement of full conversion of the reactants (H2, CO2). Despite differences in operation, both environmental technologies and trickling bed reactors for biomethanation are designed with a large surface area to facilitate: (1) the gas–liquid mass transfer of gases which are often sparingly soluble, and (2) the formation of an active microbial biofilm which can catalyze the conversion of the dissolved gases. Several studies have demonstrated the applicability of the trickling bed reactor for biomethanation (Table 12.3). Many of these studies have achieved high CH4 concentration in the product gas, and thereby demonstrated the potential of the

biomethanation process; however, the majority of the studies have been conducted in laboratory reactors of a few liters volume. Upscaling of the technology is therefore required at the current state to demonstrate the applicability of the trickling bed technology as a methanation platform for biogas upgrading at an industrially relevant scale, and on that basis compare process costs with the more mature CSTR technology.

12.8 Future perspectives and applications

Biomethanation represents a promising new technology that can be used for upgrading of biogas through the conversion of electricity and biogas-CO2 to high-grade CH4. The technology’s strengths, weaknesses, opportunities, and threats, which will define its role in a future energy system based on renewable sources, are identified in Table 12.6 (Agneessens, 2018).

Table 12.6

Strengths

• Biomethane is a flexible and mature energy carrier giving it versatile and immediate application • Use of existing gas g Opportunities

• Binding and more ambitious targets for renewable transport fuels • Use of biomethane as a building block for producti

An essential strength of all types of biogas upgrading technologies is the compliance of the produced biomethane with the globally extensive fossil natural gas infrastructure that gives it immediate application in all energy sectors (heating, electricity, transport), and renders bulk energy storage possible in the natural gas grid. A cornerstone in this CO2 conversion is that the H2 used is produced from electricity of renewable origin, and commercialization of biomethanation is consequently conditional on the continued implementation of wind turbines and solar s—a development which is ongoing in many developed countries, resulting in electricity prices that can compete with those of fossil origin (IRENA, 2019b). Although the price of electrolysis-derived H2 currently is higher than that of fossil origin, the continued development in both renewable electricity generation and electrolysis technology s the further development of the biomethanation technology. Here it is also important to note that conventional anaerobic digestion also includes significant costs for purchasing and transporting biomasses and the cost of H2 should therefore be compared to the cost of those biomasses. The use of renewable electricity for biogas upgrading through biomethane production is simultaneously expected to be a key feature in the renewable energy system by directing renewable power to the transport sector, where renewable carbon-based energy carriers are urgently needed as fuels to render sustainable transport possible (IRENA, 2018a). Being a power-to-gas technology, biomethanation-based biogas upgrading is hence ing the political focus of achieving a CO2-neutral economy. The biomethanation technology is broadly speaking still only validated at a smaller scale and it therefore still needs some development to reach commercial maturity. Although H2 is readily converted by the methanogens, the biomethanation technology is still limited by the low solubility of H2, entailing further development of the biological and technical scaffold. Due to the dependence on H2, the future deployment of the technology will also require further development of electrolysis technology to reduce hardware cost and increase the power-to-H2 efficiency to reduce energy conversion losses involved in the biomethanation process. Götz et al. (2016) estimated that 55% of the electric energy can be retained as chemical energy in power-to-CH4 conversion processes. The relatively high energy losses involved in methanation processes compared to modern battery technologies (60%–90%; Aneke and Wang, 2016), or direct usage of H2 (70%; Götz et al., 2016), could mean that the application of

the biomethane will be scenario specific and depend on regional energy infrastructures. To compete with the prices of fossil-based natural gas, technical developments have to be ed by political action to promote the attractiveness of substituting fossil natural gas for renewable biomethane, similar to existing policies for conventional biomethane production in some countries (Schmid et al., 2019). The needed actions may be to secure a sufficiently high price for the produced biomethane along with a low price of the raw materials needed for biomethanation. Reduced costs for the reactants of biomethanation can be achieved in several ways that will require legislative action to secure the right incentives, which could include: (1) reducing the electricity cost, for example, reducing grid tariffs for grid operators which constitute a large part of the electricity price today (Benjaminsson et al., 2013); (2) increasing the penalty of CO2 emission, thereby favoring the deployment of CO2 capture and conversion technologies like biomethanation. The latter could be ed by carbon taxes, making it attractive for CO2 emitters to capture and convert their CO2. Although biogas-CO2 provides a suitable and inexpensive source for biomethanation, the technology is not limited to this industry and has the potential to provide carbon capture and conversion for other stationary industrial CO2 emitters like cement plants, steel industries, and incineration plants. Biomethanation technology thus constitutes a promising technology not only for biogas upgrading and energy conversion, but also for reducing CO2 emissions in other industrial sectors.

12.9 Conclusions

Biological methanation represents part of a power-to-gas technology process chain for converting electricity to renewable methane. Biomethanation is based on using methanogenic Archaea which can utilize electricity-sourced H2 and biogas-CO2 as an electron donor and acceptor, respectively, in their metabolism to form CH4 as a gaseous end-product. The last decade has resulted in intense research on hydrogenotrophic methanogens and the factors controlling their activity, yielding both knowledge and technical solutions for biomethanation. Technologies are today developed using both complex cultures and pure cultures of methanogenic microorganisms, demonstrating technology which is both robust and flexible. There is today no consensus on the design of the technology, however, the most promising reactor designs are currently based on either CSTR technology or biofilm-based trickle-bed reactors designed for optimizing the gas–liquid mass transfer of H2 to the methanogens, which has been identified as one of the most significant challenges for the technoeconomic feasibility of biomethanation technologies. At the moment, most research and development is conducted in lab-scale reactors. However, attempts to scale up technologies are ongoing and involved industries expect proposed solutions to reach full scale in the coming years.

Abbreviations list

VFA volatile fatty acid

DIC dissolved inorganic carbon

CSTR continuous stirred tank reactor

SWOT strengths, weaknesses, opportunities, threats

References

//hornsdalepowerreserve.com.au/, 2019 //hornsdalepowerreserve.com.au/ [WWW Document], 2019. (accessed 05.01.20). Agneessens, 2018 Agneessens, L.M., 2018. The Power of the Small: Microbial Conversion of Hydrogen to Methane during Anaerobic Digestion. Aarhus University. Agneessens et al., 2017 Agneessens LM, Ottosen LDM, Voigt NV, et al. In-situ biogas upgrading with pulse H2 additions: the relevance of methanogen adaption and inorganic carbon level. Bioresour Technol. 2017;233:256–263 Available from: https://doi.org/10.1016/j.biortech.2017.02.016. Agneessens et al., 2018 Agneessens LM, Ottosen LDM, Andersen M, Berg Olesen C, Feilberg A, Kofoed MVW. Parameters affecting acetate concentrations during in-situ biological hydrogen methanation. Bioresour Technol. 2018;258:33–40 Available from: https://doi.org/10.1016/j.biortech.2018.02.102. Alitalo et al., 2015 Alitalo A, Niskanen M, Aura E. Biocatalytic methanation of hydrogen and carbon dioxide in a fixed bed reactor. Bioresour Technol. 2015;196:600–605. Al-Mamoori et al., 2017 Al-Mamoori A, Krishnamurthy A, Rownaghi AA, Rezaei F. Carbon capture and utilization update. Energy Technol. 2017;5:834– 849 Available from: https://doi.org/10.1002/ente.201600747. Andreasen et al., 2013 Andreasen RR, Liu D, Ravn S, Feilberg A, Poulsen TG. Air–water mass transfer of sparingly soluble odorous compounds in granular biofilter media. Chem Eng J. 2013;220:431–440 Available from: https://doi.org/10.1016/J.CEJ.2012.12.087. Aneke and Wang, 2016 Aneke M, Wang M. Energy storage technologies and

real life applications – a state of the art review. Appl Energy. 2016;179:350–377 Available from: https://doi.org/10.1016/J.APENERGY.2016.06.097. Angelidaki et al., 2018 Angelidaki I, Treu L, Tsapekos P, et al. Biogas upgrading and utilization: current status and perspectives. Biotechnol Adv. 2018;36:452– 466 Available from: https://doi.org/10.1016/J.BIOTECHADV.2018.01.011. Bassani et al., 2015 Bassani I, Kougias PG, Treu L, Angelidaki I. Biogas upgrading via hydrogenotrophic methanogenesis in two-stage continuous stirred tank reactors at mesophilic and thermophilic conditions. Environ Sci Technol. 2015;49:12585–12593 Available from: https://doi.org/10.1021/acs.est.5b03451. Bassani et al., 2017 Bassani I, Kougias PG, Treu L, Porté H, Campanaro S, Angelidaki I. Optimization of hydrogen dispersion in thermophilic up-flow reactors for ex situ biogas upgrading. Bioresour Technol. 2017;234:310–319 Available from: https://doi.org/10.1016/J.BIORTECH.2017.03.055. Benjaminsson et al., 2013 Benjaminsson, G., Benjaminsson, J., Rudberg, R.B., 2013. Power-to-Gas – A technical review. Malmö. Burkhardt and Busch, 2013 Burkhardt M, Busch G. Methanation of hydrogen and carbon dioxide. Appl Energy. 2013;111:74–79 Available from: https://doi.org/10.1016/J.APENERGY.2013.04.080. Burkhardt et al., 2015 Burkhardt M, Koschack T, Busch G. Biocatalytic methanation of hydrogen and carbon dioxide in an anaerobic three-phase system. Bioresour Technol. 2015;178:330–333 Available from: https://doi.org/10.1016/J.BIORTECH.2014.08.023. Chen and Wilde, 1991 Chen HC, Wilde F. The effect of dissolved oxygen and aeration rate on antibiotic production of Streptomyces fradiae. Biotechnol Bioeng. 1991;37:591–595 Available from: https://doi.org/10.1002/bit.260370615. Chen et al., 2019 Chen X, Ottosen LDM, Kofoed MVW. How low can you go: methane production of Methanobacterium congolense at low CO2 concentrations. Front Bioeng Biotechnol. 2019;7:34. Corbellini et al., 2018 Corbellini V, Kougias PG, Treu L, Bassani I, Malpei F, Angelidaki I. Hybrid biogas upgrading in a two-stage thermophilic reactor.

Energy Convers Manage. 2018;168 Available from: https://doi.org/10.1016/j.enconman.2018.04.074. Cussler, 1997 Cussler EL. Fundamentals of mass transfer. Diffusion – Mass Transfer in Fluid Systems Cambridge: Cambridge University Press; 1997;237– 273. Dannesboe et al., 2020 Dannesboe C, Hansen JB, Johannsen I. Catalytic methanation of CO2 in biogas: experimental results from a reactor at full scale. React Chem Eng. 2020;5:183–189 Available from: https://doi.org/10.1039/C9RE00351G. de Poorter et al., 2007 de Poorter LMI, Geerts WJ, Keltjens JT. Coupling of Methanothermobacter thermautotrophicus methane formation and growth in fedbatch and continuous cultures under different H2 gassing regimens. Appl Environ Microbiol. 2007;73 740 LP – 749. Available from: https://doi.org/10.1128/AEM.01885-06. De Vrieze et al., 2015 De Vrieze J, Saunders AM, He Y, et al. Ammonia and temperature determine potential clustering in the anaerobic digestion microbiome. Water Res. 2015;75:312–323 Available from: https://doi.org/10.1016/j.watres.2015.02.025. Delhoménie and Heitz, 2005 Delhoménie M-C, Heitz M. Biofiltration of air: a review. Crit Rev Biotechnol. 2005;25:53–72 Available from: https://doi.org/10.1080/07388550590935814. Dupnock and Deshusses, 2017 Dupnock TL, Deshusses MA. High-performance biogas upgrading using a biotrickling filter and hydrogenotrophic methanogens. Appl Biochem Biotechnol. 2017;183:488–502 Available from: https://doi.org/10.1007/s12010-017-2569-2. Dupnock and Deshusses, 2019 Dupnock TL, Deshusses MA. Detailed investigations of dissolved hydrogen and hydrogen mass transfer in a biotrickling filter for upgrading biogas. Bioresour Technol. 2019;290:121780. Electrochaea, 2017 Electrochaea, 2017. GmbH, Final Report: Power-to-Gas via Biological Catalysis (P2G-Biocat). Electrochaea, 2018 Electrochaea, 2018. Specifications of Electrochaea’s BioCat

methanation plants. Fenchel et al., 1988 Fenchel T, King GM, Blackburn TH. Bacterial Biogeochemistry second ed. London: Academic Press; 1988. Fortuny et al., 2011 Fortuny M, Gamisans X, Deshusses MA, Lafuente J, Casas C, Gabriel D. Operational aspects of the desulfurization process of energy gases mimics in biotrickling filters. Water Res. 2011;45:5665–5674 Available from: https://doi.org/10.1016/j.watres.2011.08.029. Garcia-Ochoa and Gomez, 2009 Garcia-Ochoa F, Gomez E. Bioreactor scale-up and oxygen transfer rate in microbial processes: an overview. Biotechnol Adv. 2009;27:153–176 Available from: https://doi.org/10.1016/j.biotechadv.2008.10.006. Garcia-Robledo et al., 2016 Garcia-Robledo E, Ottosen LDM, Voigt NV, Kofoed MW, Revsbech NP. Micro-scale H2–CO2 dynamics in a hydrogenotrophic methanogenic membrane reactor. Front Microbiol. 2016;7:1276 Available from: https://doi.org/10.3389/fmicb.2016.01276. Götz et al., 2016 Götz M, Lefebvre J, Mörs F, et al. Renewable Power-to-gas: a technological and economic review. Renew Energy. 2016;85:1371–1390 Available from: https://doi.org/10.1016/j.renene.2015.07.066. Jacob Gronholt-Pedersen Reuters, 2020 Jacob Gronholt-Pedersen Reuters, 2020. Denmark Sources Record 47% of Power From Wind in 2019. Guneratnam et al., 2016 Guneratnam AJ, Ahern E, Fitzgerald JA, et al. Study of the performance of a thermophilic biological methanation system. Bioresour Technol. 2016;225:308–315 Available from: https://doi.org/10.1016/j.biortech.2016.11.066. Holmes and Smith, 2016 Holmes DE, Smith JA. Biologically produced methane as a renewable energy source. Adv Appl Microbiol. 2016;97:1–61 Available from: https://doi.org/10.1016/BS.AAMBS.2016.09.001. IEA, 2018 IEA, 2018. Biological Methanation Demonstration Plant in Allendorf, - An Upgrading Facility for Biogas. IEA Bioenergy Task 37 Oct. 2018. IRENA, 2018a IRENA. Biogas for Road Vehicles Technology Brief Abu Dhabi:

International Renewable Energy Agency; 2018a. IRENA, 2018b IRENA. Hydrogen From Renewable Power: Technology Outlook for the Energy Transition Abu Dhabi: International Renewable Energy Agency; 2018b. IRENA, 2019a IRENA. Innovation Landscape for a Renewable-Powered Future Abu Dhabi: International Renewable Energy Agency; 2019a. IRENA, 2019b IRENA. Renewable Power Generation Costs in 2018 Abu Dhabi: International Renewable Energy Agency; 2019b. Jee et al., 1987 Jee HS, Yano T, Nishio N, Nagai S. Biomethanation of H2 and CO2 by Methanobacterium thermoautotrophicum in membrane and ceramic bioreactors. J Ferment Technol. 1987;65:413–418 Available from: https://doi.org/10.1016/0385-6380(87)90137-3. Jee et al., 1988a Jee HS, Nishio N, Nagai S. Continuous CH4 production from H2 and CO2 by Methanobacterium thermoautotrophicum in a fixed-bed reactor. J Ferment Technol. 1988a;66:235–238 Available from: https://doi.org/10.1016/0385-6380(88)90054-4. Jee et al., 1988b Jee HS, Nishio N, Nagai S. CH4 production from H2 and CO2 by Methanobacterium thermoautotrophicum cells fixed on hollow fibers. Biotechnol Lett. 1988b;10:243–248. Jensen et al., 2018 Jensen MB, Kofoed MVW, Fischer K, et al. Venturi-type injection system as a potential H2 mass transfer technology for full-scale in situ biomethanation. Appl Energy. 2018;222:840–846 Available from: https://doi.org/10.1016/j.apenergy.2018.04.034. Jensen et al., 2019 Jensen MB, Strübing D, Jonge N, et al. Stick or leave – pushing methanogens to biofilm formation for ex situ biomethanation. Bioresour Technol. 2019;291:121784. Jensen et al., 2020 Jensen, M.B., Jensen, B., Mørck Ottosen, L.D., Wegener Kofoed, M.V., 2020a. Integrating H2 injection and reactor mixing for low-cost H2 gas-liquid mass transfer in full-scale in situ biomethanation. Biochem. Eng. J., 107869. Available from: https://doi.org/10.1016/j.bej.2020.107869.

Jensen et al., 2020b Jensen, M.B., Ottosen, L.D.M., Kofoed, M.V.W., 2020b. H2 Gas-liquid mass transfer: a key element in biological power-to-gas methanation. In Preparation. Kaster et al., 2011 Kaster A-K, Moll J, Parey K, Thauer RK. Coupling of ferredoxin and heterodisulfide reduction via electron bifurcation in hydrogenotrophic methanogenic archaea. Proc Natl Acad Sci. 2011;108 2981 LP – 2986 https://doi.org/10.1073/pnas.1016761108. Kougias et al., 2017 Kougias PG, Treu L, Benavente DP, Boe K, Campanaro S, Angelidaki I. Ex-situ biogas upgrading and enhancement in different reactor systems. Bioresour Technol. 2017;225:429–437 Available from: https://doi.org/10.1016/j.biortech.2016.11.124. Lecker et al., 2017 Lecker B, Lukas I, Lemmer A, Oechsner H. Biological hydrogen methanation – a review. Bioresour Technol. 2017;245:1220–1228. Lee et al., 2012 Lee JC, Kim HJ, Chang WS, Pak D. Biological conversion of CO2 to CH4 using hydrogenotrophic methanogen in a fixed bed reactor. Chem Technol Biotechnol. 2012;87:844–847. Levén et al., 2007 Levén L, Eriksson ARB, Schnürer A. Effect of process temperature on bacterial and archaeal communities in two methanogenic bioreactors treating organic household waste. FEMS Microbiol Ecol. 2007;59:683–693. Lewandowska-Bernat and Desideri, 2017 Lewandowska-Bernat A, Desideri U. Opportunities of power-to-gas technology. Energy Procedia Elsevier Ltd. 2017;4569–4574 Available from: https://doi.org/10.1016/j.egypro.2017.03.982. Lewis and Whitman, 1924 Lewis WK, Whitman WG. Principles of gas absorption. Ind Eng Chem. 1924;16:1215–1220 Available from: https://doi.org/10.1021/ie50180a002. Liu et al., 2016 Liu R, Hao X, Wei J. Function of homoacetogenesis on the heterotrophic methane production with exogenous H2/CO2 involved. Chem Eng J. 2016;284:1196–1203 Available from: https://doi.org/10.1016/J.CEJ.2015.09.081. Lund et al., 2016 Lund H, Østergaard PA, Connolly D, et al. Energy storage and

smart energy systems. Int J Sustain Energy Plan Manage. 2016;11:3–14 Available from: https://doi.org/10.5278/ijsepm.2016.11.2. Luo and Angelidaki, 2012 Luo G, Angelidaki I. Integrated biogas upgrading and hydrogen utilization in an anaerobic reactor containing enriched hydrogenotrophic methanogenic culture. Biotechnol Bioeng. 2012;109:2729– 2736 Available from: https://doi.org/10.1002/bit.24557. Luo and Angelidaki, 2013 Luo G, Angelidaki I. Co-digestion of manure and whey for in situ biogas upgrading by the addition of H2: process performance and microbial insights. Appl Microbiol Biotechnol. 2013;97:1373–1381 Available from: https://doi.org/10.1007/s00253-012-4547-5. Luo et al., 2012 Luo G, Johansson S, Boe K, Xie L, Zhou Q, Angelidaki I. Simultaneous hydrogen utilization and in situ biogas upgrading in an anaerobic reactor. Biotechnol Bioeng. 2012;109:1088–1094 Available from: https://doi.org/10.1002/bit.24360. Madigan et al., 2019a Madigan, M.T., Bender, K.S., Buckley, D.H., Sattley, W.M., Stahl, D.A., 2019a. Chapter 21 – Nutrient cycles. In: Brock Biology of Microorganisms. pp. 687–707. Madigan et al., 2019b Madigan MT, Bender KS, Buckley DH, Sattley WM, Stahl DA. Chapter 3 – Microbial metabolism. Brock Biology of Microorganisms Pearson 2019b;109–137. Maegaard et al., 2019 Maegaard K, Garcia-Robledo E, Kofoed MVW, et al. Biogas upgrading with hydrogenotrophic methanogenic biofilms. Bioresour Technol. 2019;287:121422 Available from: https://doi.org/10.1016/J.BIORTECH.2019.121422. Martin et al., 2013 Martin, M.R., Fornero, J.J., Stark, R., Mets, L., Angenent, L.T., 2013. A single-culture bioprocess of Methanothermobacter thermautotrophicus to upgrade digester biogas by CO2 -to-CH4 conversion with H2. Archaea, 2013, Artile ID 157529. Available from: https://doi.org/10.1155/2013/157529. Mathiesen et al., 2015 Mathiesen BV, Lund H, Connolly D, et al. Smart energy systems for coherent 100% renewable energy and transport solutions. Appl Energy. 2015;145:139–154 Available from:

https://doi.org/10.1016/j.apenergy.2015.01.075. Nishimura et al., 1992 Nishimura N, Kitaura S, Mimura A, Takahara Y. Cultivation of thermophilic methanogen KN-15 on H2-CO2 under pressurized conditions. J Ferment Bioeng. 1992;73:477–480 Available from: https://doi.org/10.1016/0922-338X(92)90141-G. NIST Chemistry WebBook, 2020 NIST Chemistry WebBook, 2020. NIST Chemistry WebBook [WWW Document]. Nord Pool [WWW Document] Nord Pool [WWW Document], <www.nordpoolgroup.com>. Ottosen et al., 2011 Ottosen LDM, Juhler S, Guldberg LB, Feilberg A, Revsbech NP, Nielsen LP. Regulation of ammonia oxidation in biotrickling airfilters with high ammonium load. Chem Eng J. 2011;167:198–205 Available from: https://doi.org/10.1016/J.CEJ.2010.12.022. Peillex et al., 1990 Peillex J-P, Fardeau M-L, Belaich J-P. Growth of Methanobacterium thermoautotrophicum on H2-CO2: high CH4 productivities in continuous culture. Biomass. 1990;21:315–321 Available from: https://doi.org/10.1016/0144-4565(90)90080-4. Poehlein et al., 2012 Poehlein, A., Schmidt, S., Kaster, A., Goenrich, M., Vollmers, J., Bertsch, J., et al., 2012. An ancient pathway combining carbon dioxide fixation with the generation and utilization of a sodium ion gradient for ATP synthesis, PLoS One, 7 Available from: https://doi.org/10.1371/journal.pone.0033439. Pokorna et al., 2019 Pokorna D, Varga Z, Andreides D, Zabranska J. Adaptation of anaerobic culture to bioconversion of carbon dioxide with hydrogen to biomethane. Renew Energy. 2019;142:167–172 Available from: https://doi.org/10.1016/j.renene.2019.04.076. Poling et al., 2008 Poling, B.E., Thomson, G.H., Friend, D.G., et al., 2008. Physical and chemical data. In: Perry’s Chemical Engineers’ Handbook. The McGraw-Hill Companies. Porté et al., 2019 Porté H, Kougias PG, Alfaro N, Treu L, Campanaro S, Angelidaki I. Process performance and microbial community structure in

thermophilic trickling biofilter reactors for biogas upgrading. Sci Total Environ. 2019;655:529–538 Available from: https://doi.org/10.1016/j.scitotenv.2018.11.289. Rachbauer et al., 2016 Rachbauer L, Voitl G, Bochmann G, Fuchs W. Biological biogas upgrading capacity of a hydrogenotrophic community in a trickle-bed reactor. Appl Energy. 2016;180:483–490. Rachbauer et al., 2017 Rachbauer L, Beyer R, Bochmann G, Fuchs W. Characteristics of adapted hydrogenotrophic community during biomethanation. Sci Total Environ. 2017;595:912–919. Rittmann et al., 2012 Rittmann S, Seifert A, Herwig C. Quantitative analysis of media dilution rate effects on Methanothermobacter marburgensis grown in continuous culture on H2 and CO2. Biomass Bioenergy. 2012;36:293–301 Available from: https://doi.org/10.1016/j.biombioe.2011.10.038. Savvas et al., 2017 Savvas S, Donnelly J, Patterson T, Chong ZS, Esteves SR. Biological methanation of CO2 in a novel biofilm plug-flow reactor: a high rate and low parasitic energy process. Appl Energy. 2017;202:238–247 Available from: https://doi.org/10.1016/J.APENERGY.2017.05.134. Scarlat et al., 2018 Scarlat N, Dallemand J-F, Fahl F. Biogas: developments and perspectives in Europe. Renew Energy. 2018;129:457–472 Available from: https://doi.org/10.1016/J.RENENE.2018.03.006. Scherer et al., 1983 Scherer P, Lippert H, Wolff G. Composition of the major elements and trace elements of 10 methanogenic bacteria determined by inductively coupled plasma emission spectrometry. Biol Trace Elem Res. 1983;5:149–163 Available from: https://doi.org/10.1007/BF02916619. Schill et al., 1996 Schill N, Gulik WMV, Voisard D, Stockar UV. Continuous cultures limited by a gaseous substrate: development of a simple, unstructured mathematical model and experimental verification with Methanobacterium thermoautotrophicum. Biotechnol Bioeng. 1996;51:645–658. Schink, 1997 Schink B. Energetics of syntrophic cooperation in methanogenic degradation. Microbiol Mol Biol Rev. 1997;61:262–280. Schmid et al., 2019 Schmid C, Horschig T, Pfeiffer A, Szarka N, Thrän D.

Biogas upgrading: a review of national biomethane strategies and policies in selected countries. Energies. 2019;12:3803 Available from: https://doi.org/10.3390/en12193803. Seifert et al., 2014 Seifert AH, Rittmann S, Herwig C. Analysis of process related factors to increase volumetric productivity and quality of biomethane with Methanothermobacter marburgensis. Appl Energy. 2014;132:155–162 Available from: https://doi.org/10.1016/J.APENERGY.2014.07.002. Sieborg et al., 2020 Sieborg MU, Jønson BD, Ashraf MT, Yde L, Triolo JM. Biomethanation in a thermophilic biotrickling filter using cattle manure as nutrient media. Bioresour Technol Rep. 2020;9:100391 Available from: https://doi.org/10.1016/j.biteb.2020.100391. Singh et al., 2015 Singh S, Jain S, PS V, et al. Hydrogen: a sustainable fuel for future of the transport sector. Renew Sustain Energy Rev. 2015;51:623–633 Available from: https://doi.org/10.1016/j.rser.2015.06.040. Sommer and Husted, 1995 Sommer SG, Husted S. The chemical buffer system in raw and digested animal slurry. J Agric Sci. 1995;124:45–53 Available from: https://doi.org/10.1017/S0021859600071239. Sovacool, 2013 Sovacool BK. Energy policymaking in Denmark: implications for global energy security and sustainability. Energy Policy. 2013;61:829–839. Strübing et al., 2017 Strübing D, Huber B, Lebuhn M, Drewes JE, Koch K. High performance biological methanation in a thermophilic anaerobic trickle bed reactor. Bioresour Technol. 2017;245:1176–1183. Sun et al., 2015 Sun Q, Li H, Yan J, Liu L, Yu Z, Yu X. Selection of appropriate biogas upgrading technology – a review of biogas cleaning, upgrading and utilization. Renew Sustain Energy Rev. 2015;51:521–532. Sveinbjörnsson and Münster, 2017 Sveinbjörnsson, D., Münster, E., 2017. Gas conditioning and grid operation: upgrading of biogas to biomethane with the addition of hydrogen from electrolysis – futuregas project [WWW Document]. . Thema et al., 2019 Thema M, Bauer F, Sterner M. Power-to-gas: electrolysis and

methanation status review. Renew Sustain Energy Rev. 2019;112:775–787 Available from: https://doi.org/10.1016/J.RSER.2019.06.030. Trček et al., 2015 Trček J, Mira NP, Jarboe LR. Adaptation and tolerance of bacteria against acetic acid. Appl Microbiol Biotechnol. 2015;99:6215–6229 Available from: https://doi.org/10.1007/s00253-015-6762-3. Van’t Riet, 1979 Van’t Riet K. Review of measuring methods and results in nonviscous gas-liquid mass transfer in stirred vessels. Ind Eng Chem Process Des Dev. 1979;18:357–364 Available from: https://doi.org/10.1021/i260071a001. Wahid et al., 2019 Wahid R, Girma Mulat D, Christian Gaby J, Jarle Horn S. Effects of H2:CO2 ratio and H2 supply fluctuation on methane content and microbial community composition during in-situ biological biogas upgrading. Biotechnol Biofuels. 2019;12:104 Available from: https://doi.org/10.1186/s13068-019-1443-6. Weiland, 2010 Weiland P. Biogas production: current state and perspectives. Appl Microbiol Biotechnol. 2010;85:849–860 Available from: https://doi.org/10.1007/s00253-009-2246-7. Zabranska and Pokorna, 2018 Zabranska J, Pokorna D. Bioconversion of carbon dioxide to methane using hydrogen and hydrogenotrophic methanogens. Biotechnol Adv. 2018;36:707–720 Available from: https://doi.org/10.1016/j.biotechadv.2017.12.003. Zhu et al., 2019 Zhu X, Chen L, Chen Y, Cao Q, Liu X, Li D. Differences of methanogenesis between mesophilic and thermophilic in situ biogas-upgrading systems by hydrogen addition. J Ind Microbiol Biotechnol. 2019; Available from: https://doi.org/10.1007/s10295-019-02219-w.

Chapter 13

Bioelectrochemical systems for biogas upgrading and biomethane production

Nabin Aryal¹, Lars Ditlev Mørck Ottosen¹, Anders Bentien¹, Deepak Pant² and Michael Vedel Wegener Kofoed¹, ¹1Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark, ²2Separation and Conversion Technology, Flemish Institute for Technological Research (VITO), Mol, Belgium

Abstract

Biogas upgrading by employment of bioelectrochemical systems (BESs) is an emerging approach for electricity-based production of biomethane. Recent advances within the field have successfully demonstrated BES for biogas upgrading at laboratory scale under different configurations and operating conditions: in situ, ex situ, batch mode, and continuous mode. This chapter summarizes the development and status of bioelectrochemical biogas upgrading, and includes examples of multifunctional systems combining biogas upgrading with resource recovery. Insights are given into proposed electron transfer mechanisms and reported BES designs for CO2 reduction to methane. BES technology for biogas upgrading has primarily been developed to lab-scale and still has to be further developed to evaluate the current economic perspectives of the technology compared to conventional biogas-upgrading technologies.

Keywords

Biogas; biogas upgrading; bioelectrochemical systems; BES; electromethanogens; economic insight

Chapter outline

Outline

13.1 Background 363

13.2 Fundamentals of bioelectrochemical biogas upgrading 364

13.3 Methane enrichment of biogas 368

13.3.1 Electron transfer mechanism 368 13.3.2 Microbial communities in biocathode for methane enrichment 370 13.3.3 State-of-the-art bioelectrochemical biogas upgrading 370

13.4 Economical insights 375

13.5 Prospective and challenges 376

13.6 Conclusion 378

References 378

13.1 Background

Anaerobic digestion (AD) technology combines waste treatment and energy generation by producing biogas rich in methane (CH4) from different waste materials such as sludge, municipal solid waste, and agricultural residues. The produced CH4 is a versatile energy carrier that can be used as a fuel source in households, industry, and the transportation sector. Biogas is a mixture of different gases: CH4 (40%–75%), carbon dioxide (CO2) (25%–60%), and trace amounts of other gases including hydrogen (H2), hydrogen sulfide (H2S), nitrogen (N2), ammonia (NH3), carbon monoxide (CO), oxygen (O2), and siloxanes depending on feedstock and operating conditions (Aryal and Kvist, 2018). These non-CH4 constituents reduce the heating value of biogas, cause salt accumulation on processing equipment, and their emission can be hazardous to human health. Special concerns include H2S and siloxanes, which must be removed prior to downstream use due to their corrosive nature (Angelidaki et al., 2018). To upgrade the biogas to natural gas quality, the concentration of CO2 has to be decreased to increase the heating value of biogas. Various biogas-upgrading technologies like water scrubbing, chemical adsorption, membrane separation, pressure swing absorption (PSA), cryogenic separation, and H2-based chemical and biological methanation technologies are being adopted for biogas upgrading (Angelidaki et al., 2018). Water scrubbing is the most widely adopted technology for biogas upgrading and exploits that CO2 has a significantly higher aqueous solubility than CH4 to separate the two constituents like water scrubbing (Kvist and Aryal, 2019). Nonetheless, conventional biogas upgrading requires significant investment costs, is energy demanding, includes the risk of CH4 leakage, and typically emits the separated CO2 to the atmosphere (Angelidaki et al., 2018; Kvist and Aryal, 2019).

13.2 Fundamentals of bioelectrochemical biogas upgrading

As an alternative to emitting the CO2 to the atmosphere, bioelectrochemical systems (BESs) represent a technology platform for upgrading biogas through the bioelectrochemical reduction of CO2 to CH4 (Xu et al., 2014). The technology involves electrode-based reactions, mainly water splitting and oxygen evolution at the anode, with simultaneous reduction of CO2 to CH4 at the cathode. The BES is powered by an external electricity supply and catalyzed by biological agents. In BES, the active methanogenic microbial community catalyzing the reduction of CO2 to CH4 is positioned at the cathode and receives the reducing equivalents to catalyze the biological conversion of CO2 from biogas to CH4 (Cheng et al., 2009). CO2 has a redox state of +IV, whereas CH4 has a redox state of −IV and eight electrons therefore have to be transferred for the complete reduction of CO2 to CH4. The anode and cathode are separated by an ion-exchange membrane, which allows the transport of ions as shown in Fig. 13.1 and Eqs. 13.1, 13.2, and 13.3, respectively. The membrane enables protons or H2 produced from the splitting of water at the anode to be transported to the cathodic side, where they are used for the microbial reduction of CO2 to CH4, together with the externally supplied electrons. The BESs perform cathodic CO2 reduction (as shown in Eqs. 13.1 and 13.2) by use of electricity and anodic water oxidation (as shown in Eq. 13.3) (Rojas et al., 2018a,b), and thereby function as a tool for transforming electricity and CO2 to an energy carrier like CH4.

Figure 13.1 Bioelectrochemical systems used for either electricity production or methane production. (A) A microbial fuel cell (MFC) where electricity is produced through the microbial oxidation of organic matter; (B) microbial electrosynthesis (MES) where CO2 from biogas or other CO2 sources is reduced to CH4 (or other reduced compounds) by use of electrons supplied from an external power source; and (C) a microbial electrolysis cell (MEC) that generates CH4 from CO2 at the cathode by receiving electrons recovered from anodic oxidation of organic matters, which in turn is stimulated by applying an external power source. Arrows denote the proton flux across the membrane “M.”

The processes follow the reaction schemes presented below: Cathodic reduction

(13.1)

(13.2)

Anodic oxidation

(13.3)

BES combines electrochemistry with biological processes and includes different configurations: (1) microbial fuel cell (MFC), where electrons are harvested at the anode from the oxidation of organic materials and supplied to the cathode for a reduction process. However, the intention of an MFC configuration is to oxidize organic material into CO2 and electrical power, whereas the intention with microbial electrosynthesis (MES) biogas upgrading is exactly the opposite, namely to use electricity to reduce CO2 into CH4 or organics, (2) MES, where CO2 from, for example, biogas, has been utilized to produce either CH4 or chemicals, (3) microbial electrolysis cell (MEC) that generates H2 or CH4 at the cathode and organic material oxidation at the anode by external electrical power (Schröder et al., 2015). In this chapter, the broad term of BES will used for describing the the different configurations for converting biogas CO2 (Nevin et al., 2010). The bioelectrochemical route of CH4 production through electrochemical reduction of CO2 represents a promising power-to-gas technology which enables the capture and utilization of waste CO2. The produced CH4 can be injected and stored in the existing natural gas grid infrastructure, which in particular in the EU is well developed and connected across large regions. Using renewable electricity for producing CH4 enables conversion and subsequent large-scale storage of electricity using CH4 as an energy carrier. There have been several studies employing BES in combination with biogas production to reduce CO2 and enrich the CH4 concentration of the biogas. Biogas upgrading has been demonstrated with the use of BES in different studies using different configurations and at different efficiencies (Geppert et al., 2016). An important parameter to consider when evaluating the feasibility of BES is the efficiency of the electrochemical system, which can be described by the Coulombic efficiency (CE). The CE can be calculated as described in Eq. 13.4.

(13.4)

where F is Faraday constant (96,485 C/mol), C is the molar concentration of the product, V is the volume of the reaction chamber, z is the number of electrons

required to produce CH4, and I is the electrical current integrated over time ( - ) to give the total number of electrons supplied to the system. Different factors negatively affect the CE, but can overall be divided into two. The first includes internal short circuits where the electrons are conducted, or transported via redox-active molecules (redox shuttles) from anode to cathode or vice versa, where it is oxidized without taking part in the overall reduction of CO2. The magnitude of internal short circuits is heavily dependent on the design of the BES. In most cases, the anode and cathode are separated by an ion-selective membrane that effectively reduces short circuit currents. In this case it is in practice diffusion-limited and only small neutral or small molecules with either positive charge (in cation exchange membranes) or negative (in anion exchange membranes) can be transported. The second mechanism that can degrade the CE is electrochemical side reactions, where the reducing equivalents are used in reduction processes other than that involving CO2 to CH4. Again this could be reduction of redox-active molecules that does not take part in the overall CO2 reduction. An example of this is H2 evolution, assuming that the H2 at a later stage is not involved in the CO2 reduction. In addition to the CE, which is related to mass balances in the BES, an equally important parameter is the energy efficiency (EE) of the conversion, which can be expressed by

(13.5)

where is the theoretical thermodynamic Gibbs free energy for the reduction reactions (Eqs. 13.1–13.3) and is the number of moles of CO2 that are reduced during the reaction time ( ). depends on the state (pressure, temperature, and gas/liquid) of the reactants and products in Eqs. 13.1–13.3. I is the electrical current and is the voltage applied to the electrochemical system. The product between I and gives the electrical power which is integrated over the reaction time giving the total supplied energy. There are two fundamental mechanisms that reduce EE from the ideal 100%. The first is a nonideal CE (<100%) where electrons are lost due to short circuit

or electrochemical side reactions as described in the paragraph above. The second is related to energy losses in the reaction/process that are converted into heat. A formal way of describing this is by

(13.6)

where is the theoretical minimum electric potential needed for driving the reduction reaction (Eqs. 13.1–13.3). is the activation overpotential needed for initiating the reactions and is typically associated with the electrode materials and their catalytic characteristics (Eq. 13.1). is the extra total overpotential for actually driving the reaction, while is the total electrical resistance that can be divided into several contributions. is related to ohmic losses associated with electrical resistance in the electrodes or resistance of the conduction of ions through the ion-exchange membrane. is related to mass transfer/concentration polarization, that is, discharge of reduced species CH4 from the electrode that limits the availability of CO2 on the surface of the electrode. Meanwhile and are the charge transfer resistances, related to the anode and cathode, respectively, and describe losses in the electrochemical reactions.

, and can to a good approximation be assumed to be independent of the current, while the mass transfer/concentration polarization term ( ) has a strong dependence on the electrical current, in particular when the mass transfer limit is approached. In the case of BES, where typical current densities (electrical current per unit electrode area) are below 20 mA/cm² and in many cases also includes stirring/convection, normally has insignificant contributions. There are several experimental methods for probing the magnitude of the different contributions to the losses in Eq. 13.6; the most common is electrochemical impedance spectroscopy and measurement of the polarization curve. As indicated by Eq. 13.6, the potential (U) increases with current, meaning that the efficiency of the conversion decreases with increasing current (reaction rate). For the CO2 reduction process this has the implication that operating the BES with high currents will increase the energy consumption per unit CH4 produced (increased operational cost). Operating with low currents increases the EE, but requires a larger BES (increased system cost). The optimal operating current is consequently a trade-off. Furthermore, the energy efficiency is a complex function of many parameters: cell-design ( ), electrolyte and membrane ( ), electrodes and catalyst ( ), microbes ( ), and stirring/convection ( ). An overall

increase in the EE is therefore a matter of identifying the that have the highest contributions to the total resistance that subsequently can be optimized.

13.3 Methane enrichment of biogas

13.3.1 Electron transfer mechanism

Electromethanogens are assumed to be able to produce methane from CO2 utilizing electrons directly from the cathode and are therefore of special interest in BES/MEC-based biogas upgrading. The term electromethanogenesis was coined by Cheng et al. (2009) based on experimental results where direct electron uptake was observed by methanogens from a poised graphite electrode. Examples of electromethanogens include of Methanosaeta, which has shown the capability to accept electrons directly for the CO2 reduction to methane when tested in BES (Blasco-Gómez et al., 2017). The electron transfer mechanisms from electrodes to electromethanogens are not fully elucidated and still highly debated (Tremblay et al., 2017; Gahlot et al., 2020). The proposed mechanisms of cathodic electron transfer are either as direct electron transfer or through mediator-dependent electron transfer (also called indirect electron transfer). In direct electron transfer, the microorganisms are capable of directly accepting the electrons from the cathode. Direct electron transfer requires physical between the solid electrode and extracellular components of microbes or with an associated cofactor or redox-active enzyme (Reguera et al., 2005; Rotaru et al., 2014). Extracellular conductive nanowires or c-cytochrome on the outer membrane extension have also been suggested to be involved in electron transportation for the reduction of CO2 (Tremblay et al., 2017). Methanogens can also receive the required reducing equivalents through indirect electron transfer, where electron transfer is mediated by different redox-active molecule. Mediators such as formate, H2, ammonia, and Fe(II) have been proposed to mediate the electron transfer from the cathode to the microorganism (Liu et al., 2012; Tremblay et al., 2017). Soluble redox mediators like riboflavin, and phenazine hydrogenases released from active or lysed cells are likely to be involved in electron transfer (Doud and Angenent, 2014; Morita et al., 2011). Batlle-Vilanova et al. experimentally demonstrated that methane could be

produced, either directly via direct electron transfer or indirectly via hydrogenmediated hydrogenotrophic methanogenesis, while upgrading biogas (BatlleVilanova et al., 2015), as shown in Fig. 13.2. Hydrogenotrophic methanogenesis, where methane is produced indirectly with H2 as the mediator, is often reported to be the dominating process when the cathode potential is above –400 mV versus an Ag/AgCl reference electrode (Blasco-Gómez et al., 2017). Hydrogen production occurs electrochemically but has been shown to be enhanced by the presence of electrode-associated microorganisms (Jourdin et al., 2016b).

Figure 13.2 Bioelectrochemical system for biogas upgrading where “M” is an ion-exchange membrane: (A) direct or indirect and mediated electron transfer mechanism involved in BES. “Med” is an electron mediator. (B) Surface-modified electrode-equipped BES reactor where modification leads to increase the surface for reaction. The box outside represents the surface modification lead microporous structure developed that enhances the biofilm formation.

The type of electrode materials used can also impact the electron transfer since cathode materials influence the microbial interaction in electron transfer for CO2 reduction from biogas (Park et al., 2018; Zhang et al., 2012). The electrode material and design are furthermore important for the microbial interaction and overall performance of the BES. It has been reported that cathode materials should possess beneficial properties, in particular (1) high conductivity to reduce ohmic loss, (2) biocompatibility to allow microbial colonization on the electrode surface, (3) high surface area, and (4) low cost (Aryal et al., 2017; Hou et al., 2013). An electrode with a high surface area will allow a higher number of microorganisms to colonize the surface compared to standard plate electrodes. Electrodes with a high surface area have therefore been developed to increase the electrochemically active surface area (Fig. 13.2B). An example of this is the microporous electrode of multiwalled carbon nanotube reticulated vitreous carbon (MWCNT-RVC) with a 0.6 mm pore size, which has been shown to be suitable for the removal of CO2 for the production of acetate from synthetic biogas (Jourdin et al., 2016a).

13.3.2 Microbial communities in biocathode for methane enrichment

The electron source for microbial biogas upgrading is supplied through the electrode material and will favor a close association between microorganisms and the electrode based on the mechanisms of electron transfer outlined above

(Batlle-Vilanova et al., 2015; Bo et al., 2014). The microbial communities developing at the cathodes depend on several factors. Especially, the cathode potential has been found to impact the community composition of the electromethanogens at the cathode. The application of a negative potential associated with H2 production has not surprisingly been found to favor the growth of hydrogenotrophic methanogens (Cerrillo et al., 2017; Gajaraj et al., 2017), and has been observed for reactors operated in both continuous and batch modes for biogas upgrading (Bo et al., 2014). Cerrillo et al. (2017) concurrently found that methanogens of the hydrogenotrophic family Methanobacteriaceae (predominately Methanobrevibacter genus) dominated after 95 days of operation with similar current density and methane production rate, independently of the origin of the microbial inoculum (Cerrillo et al., 2017).

13.3.3 State-of-the-art bioelectrochemical biogas upgrading

BES for converting CO2 to CH4 can be operated with different reactor configurations and operating procedures. The recent state-of-the-art for biogas upgrading in laboratory-scale BES systems is presented in Table 13.1. One important configuration of the system is whether the system is configured as in situ or ex situ, which refers to how the CO2 is supplied to the methanation reaction. In in situ systems, the cathodic chamber contains organic material that is degraded through anaerobic processes like in a conventional anaerobic digester. The degradation processes produce CO2, which can then be used for bioelectrochemical methanation in the same chamber. In ex situ systems, the CO2 is supplied from an external source to the cathodic chamber, where a specialized community of methanogens catalyzes the CO2 reduction to methane. The added CO2 can be either in the form of pure CO2 or part of raw biogas.

Table 13.1

Electrode material

Operating mode

Reactor design

Graphite

Batch in situ

H cell

Batch ex situ

Biogas from AD fed to H cell

0.4 A/m²

Continuous in situ

Single-chamber

3 A/m²

Stainless steel of reactor acts as electrode

Batch in situ

Coupled MEC-AD (single-cham

Graphite block

Batch in situ

2 chamber BES

Continuous in situ

0.2 A/m²

Three fold faster in CH4 produc

MWCNT-RVC

Batch

H cell

Stainless steel of reactor as electrode

Batch in situ

Barrel-shaped single-chamber B

RVC

Batch in situ

Single-chamber BES

Carbon felt

Batch in situ

Two-chamber BES

Batch in situ

Single-chamber BES

–0.002 A

Titanium woven wire mesh coated with Pta

Batch in situ

MESC

Graphite granular

Batch in situ

Three-compartment BES

Cu-Ni- and Fe-coated graphite carbon

Batch in situ

AD-MEC single-chamber reacto

Stainless steel woven mesha

Batch in situ

MESC

Graphite granular

Batch in situ

Two-sided cathode compartmen

0.13 A

49% CO2 removal

108±15

Graphite granulara

Batch in situ

2 chamber BES

Carbon fiber brush

Batch in situ

Single-chamber BES

Graphite plate

Continuous in situ

Continuous flow MEC reactor

Graphite plate

Continuous in situ

H cell

AD, Anaerobic digestion; BES, bioelectrochemical system; Cu, copper; Fe, iron; FWTP, food waste treatment plant; MEC, microbial electrochemical cell; MESC, microbial electrolytic capture separation and regeneration cell; MFC, microbial fuel cell; MSW, municipal solid waste; MWCNT, multiwalled carbon nanotube; MWWTP, municipal waste water treatment plant; Ng, not given; Ni, nickel; Pt, platinum; RVC, reticulated vitreous carbon; WWTP, wastewater treatment plant; UASB, upflow anaerobic sludge blanket digestion.

aMaximum value calculated.

In a proof-of-concept study on methane enrichment using BES, Xu et al. (2014) tested the BES in both continuous and batch mode for in situ (integrating the BES within the digester), and ex situ (coupling the BES as an external unit downstream of the biogas reactor) biogas upgrading. Xu et al. (2014) reported that both in situ and ex situ systems could be employed for the conversion of biogas CO2, at current densities of 1 A/m² (in situ) and 0.4 A/m² (ex situ) (Xu et al., 2014). Following this proof-of-concept study, several research groups have developed different designs for BES biogas upgrading. These designs have included both single-chamber systems where the cathode and anode are not separated by a membrane (membraneless) and designs with two or more compartments (multichamber). Both single-chamber and dual-chamber BES have been combined with AD (Liu et al., 2017). However, the electrochemical reactions at the anode produce O2, which is toxic to the obligate anaerobic methanogens. O2 formation is a major disadvantage of a single-chamber system, while the membrane limits oxygen diffusion from the anode to cathode in multichamber systems. Multichamber systems are therefore considered as the most promising system for BES upscaling (Krieg et al., 2014). Using a dualchamber BES, Liu et al. (2017) showed that 98%–100% methane enrichment could be obtained. Recent designs in BES cell design have spawned the development of more complex systems, trying to combine biogas upgrading with other functionalities. Recently, a multichamber (more than two compartments) microbial electrochemical separation cell (MESC) was used for simultaneous

biogas upgrading and recovery of ions such as NH4+, OH−, HCO3−, and CH3COO− (Zeppilli et al., 2017, 2019). Another study applied a fourcompartment BES system with an additional two-chamber system for CO2 absorption and CO2 regeneration (where biology was not involved) apart from the anode and biocathode compartments (where biology was involved) to improve the methane enrichment by separating the CO2 (Kokkoli et al., 2018). The same research group furthermore tested a three-compartment BES system for simultaneous anodic wastewater oxidation (where biology was involved), biogas upgrading at the cathode (where biology was involved), and a third compartment used for CO2 sorption and regeneration (where biology was not involved) (Jin et al., 2017). According to the authors this multichamber BES system had several advantages over the single- and dual-compartment systems with potential to replace the traditional biogas upgrading systems because: (1) necessary chemicals such as acid and alkaline could be generated in situ; (2) the methane loss was as low as 1.4% while upgrading; (3) and the separated CO2 could be utilized for other industrial applications (Jin et al., 2017). Although these multicompartment systems are still in their infancy, they demonstrate the potential for developing multipurpose BES systems for simultaneous biogas upgrading and recovery of other chemicals.

13.4 Economical insights

A mature BES-based biogas-upgrading technology can be expected to be benchmarked against current commercial biogas-upgrading technologies such as water scrubber, amine, PSA, and membrane-based systems. The cost of biogas upgrading with conventional technologies is inversely related to the plant size (Sun et al., 2015), and typically ranges between 0.12–0.15 €/m³ methane with a plant capacity of 1000 m³/h raw biogas (IRENA, 2018). However, BES systems cannot be directly compared to the conventional upgrading system, since the BES systems valorize the CO2, which is otherwise emitted to the atmosphere by the conventional upgrading technologies. The capture and reuse of CO2 could be a significant selling point of BES-based biogas upgrading in the context of reducing the anthropogenic emissions of CO2. The BES systems should therefore be evaluated against other CO2-valorizing biogas-upgrading technologies such as biological and chemical methanation, where CO2 is upgraded through the addition of exogenous H2. Benjaminsson et al. (2013) compared the cost of two chemical methanation systems as well as two earlystage biological methanation systems for upgrading biogas to natural gas quality. The biological methanation systems were found to have a lower cost (0.118– 0.127 €/kWh CH4), compared to the chemical catalysis methanation systems (0.130–0.159 €/kWh CH4) (Benjaminsson et al., 2013). However, the methane production cost was found to be strongly dependent on the size of the systems. Future cost-estimation and profitability of biogas upgrading using BES should be benchmarked against chemical and biological methanation. The development of BES is primarily restricted to lab-scale systems, and the economic assessment of BES technology for biogas upgrading has to this date been limited. Nonetheless, one of the main bottlenecks for commercial exploitation of BES is the very low current density that would result in rather large and costly electrochemical cells (stacks). To date, the highest reported current densities for BES are about 20 mA/cm² (Jourdin et al., 2016a,b). For comparison, the similar stacks/electrochemical cells for water electrolysis or flow batteries typically operate with maximum current densities in the range >500 mA/cm² (Buttler and Spliethoff, 2018; Huskinson et al., 2014; Lin et al.,

2015). Typical costs for water electrolysis or flow battery stacks are about 0.1 €/cm² (stack unit alone). A BES system operating with ideal 100% CE would require 8.6 kA to reduce CO2 to CH4 at a rate of 1 m³/h. With optimistic current densities of 20 mA/cm² this will result in a BES stack with an internal area of about 43 m² and a corresponding cost of about 43,000 €. If it is optimistically assumed that the stack is operated 24 h a day and depreciated over 5 years, this will result in a cost of about 1 €/m³ or 0.1 €/kWh of CO2 reduced to CH4. This cost only includes the depreciation of the stack and not additional costs related to plant construction or operational expenditures (electricity, maintenance, etc.) and more realistic numbers are most likely one order of magnitude higher. The economic challenge of deploying BES for biogas upgrading in large scale is further illustrated using the 1200 m³ biogas reactor at the research facility Foulum at Aarhus University as an example. The biogas production rate from the 1200 m³ manure-based reactor with an active volume of 1100 m³ is on average 4800 m³ per day. The biogas has a methane concentration of 55% by volume and the CO2 production is consequently 2160 m³ per day. One of the highest BES rates reported is 12.5 LCH/L/d, using a double-chambered BES at current densities up to 35 A/m² and at 67% CE (Geppert et al., 2016). The cathodic chamber had a volumetric size of 84.5 cm³ with a projected electrode surface area of 169 cm², and purified CO2 was utilized instead of biogas (Geppert et al., 2019). Assuming 1:1 upscaling is possible, a cathodic chamber volume of 173 m³ would be required for converting the produced CO2 from the 1200 m³ reactor. This translates into a required electrode surface area of 34,396 m² for upgrading 2160 m³ CO2 per day. With an estimated cost of 0.1 €/cm² the stack/electrode cost alone would exceed 34 M€. The analysis clearly underlines that the single most important challenge for developing BES into a mature technology is the low current density that has to be improved by at least one or two orders of magnitude and have stable operation for years. The BES at its current state therefore still requires significant development to reach industrial maturity, especially with regard to optimization of CH4 production rate.

13.5 Prospective and challenges

Biogas upgrading using BES has gained intensive interest as an approach to convert electrical energy into fuel while cleaning the biogas. The technological development of BES-based biogas upgrading still needs to be optimized at the laboratory scale before the technology can be up-scaled and tested at the relevant technical and commercial scale. Increased efficiency of a BES will automatically lead to increased current density as they are fully coupled. As earlier outlined the low current density in the BES system is the main obstacle for exploitation of BES technology because of the consequent high cost of the electrochemical cells/stacks. Thus, cost reductions is to a large degree primarily a matter of identifying routes for increasing the current density/efficiency. BES combines both microbiology and electrochemistry, and new knowledge to engineer the electrode and reactor configuration to enhance the electron transfer mechanism to the microorganisms should therefore constitute an important part of future research (Aryal et al., 2019; Blasco-Gómez et al., 2017). The bioelectrochemical biogas upgrading performance is significantly dependent on the biocatalytic activity, redox condition, and operational parameter of the reactor. Thus, further development of BES requires the combined expertise from other scientific disciplines such as electrochemistry, microbiology, and materials sciences. The losses at the electrodes, ohmic losses, charge transfer resistance, and pH gradient collectively determine the overall internal resistance of the BES reactor. Further research on how to control and reduce the losses will be beneficial for scaling up the technology and could pave the way for reaching higher current densities needed to obtain a compact and cost-effective technology. Optimizing with respect to the high columbic efficiency and low overpotential in the BES reactor systems, as well as achieving steady-state high current densities, are the main challenges for improving the technical and economic feasibility of using bioelectrochemical systems for biogas upgrading. An important parameter in increasing the columbic efficiency is an effective charge transfer from cathode to active methanogens, which can be improved by optimizing the surface between electrodes and organisms. Therefore research has to focus on improving the

biocatalysis rate by (1) reducing the electrochemical losses through optimizing BES designs; (2) selecting electroactive species of microbes and establishing highly active biosurfaces; and (3) on reactor design to ensure an effective supply of CO2 that will allow a high rate of uptake of CO2. Highly active species could either be selected from enrichment of naturally occurring methanogens or alternatively by metabolically engineering species. Fundamental research focusing on elucidating the electron transfer mechanism in microbial electrocatalysis, which has been inconclusively studied among the methanogens, could be an important part of identifying and optimizing this association between microbe and electrode. In addition to demonstrating the improved efficiency of single elements on a labscale it is equally important to upscale and combine these learnings into stacks/electrochemical cells in order to test the feasibility of the technology. The importance of up-scaling is underlined by a comparable development within electrolysis and flow batteries, which has reduced the typical stack cost to 0.1 €/cm²—a development which has only been possible because of the large efforts related to optimization and demonstration of these optimizations on a larger scale. Improvements within cell-design and microbe–electrode interactions could also be used for BES technologies with other purposes than CH4 formation. The coupling of BES technology and fermentation technology has been proposed to constitute an economically promising way of utilizing electrical energy for the production of commodity chemicals like acetate from anaerobic fermentation processes (Christodoulou et al., 2017). CO2 from biogas could here represent a promising source of carbon in electrochemically assisted microbial fermentation and at the same time upgrade the removal of the CO2 from the biogas. A recent study on the coupling of AD and bioelectrochemical system has shown the potential economic feasibility of commercial applications (Christodoulou and Velasquez-Orta, 2016). New fermentation platforms to produce high-value products from methane fermentation such as ectoine, sucrose, biofuels, biopolymers, metal chelating proteins, enzymes, and/or heterologous proteins by a methanotrophic biocatalysis process have also shown some promise and could potentially be combined with CO2 supplied from biogas (Blasco-Gómez et al., 2017; Christodoulou et al., 2017). Thus, product diversification of the bioelectrochemical reduction processes could lead to the formation of other industrially relevant applications of BES.

Combining biogas upgrading with other functions like resource recovery or wastewater treatment could furthermore pave the way for BES implementation. EcoVolt and BioVolt Reactors from Cambrian Innovation Inc. are already commercialized as technologies to harvest renewable methane gas and electricity while treating industrial wastewater (Aryal et al., 2018; Blasco-Gómez et al., 2017). These results demonstrate the potential development of BES for commercial applications.

13.6 Conclusion

This chapter provides an overview of the research advances within biogas upgrading using BES, which combines electrochemistry and microbiology. Methanogenic microorganisms form the basis for this upgrading where biogas CO2 is reduced to CH4, thereby enriching the CH4 concentration of the biogas. Questions on how reducing equivalents are supplied from the cathode to the microorganisms remain the subject of intense research and methanogensis involving both direct electron transfer (electromethanogenesis) and indirect electron transfer through H2 has been reported. Regardless of the exact mechanism, several studies have demonstrated that the CH4 content of biogas could be enriched using different reactor configurations, both in situ and ex situ, and batch and continuous mode. These observations have led to the development and optimization of different BES configurations for biogas upgrading, including multicompartment BES reactors and new electrode designs. Methane enrichment up to 100% has been reported in the laboratory scale. However, the technological development of BES-based biogas upgrading still needs to be dramatically optimized at the laboratory scale, especially regarding the obtainable current densities, before the technology can be up-scaled and tested at the relevant technical and commercial scale.

Acknowledgments

This work was ed by Innovation Fund, Denmark – Innovationsfonden with a grant from FutureGas (IFD 5160-00006A) project and Apple Inc. as part of the APPLAUSE bio-energy collaboration with Aarhus University.

References

Angelidaki et al., 2018 Angelidaki I, Treu L, Tsapekos P, et al. Biogas upgrading and utilization: current status and perspectives. Biotechnol Adv. 2018;36:452– 466 https://doi.org/10.1016/J.BIOTECHADV.2018.01.011. Aryal et al., 2017 Aryal N, Ammam F, Patil SA, Pant D. An overview of cathode materials for microbial electrosynthesis of chemicals from carbon dioxide. Green Chem. 2017;19:5748–5760 https://doi.org/10.1039/C7GC01801K. Aryal and Kvist, 2018 Aryal N, Kvist T. Alternative of biogas injection into the Danish gas grid system—a study from demand perspective. ChemEngineering. 2018;2:43 https://doi.org/10.3390/chemengineering2030043. Aryal et al., 2018 Aryal N, Kvist T, Ammam F, Pant D, Ottosen LDM. An overview of microbial biogas enrichment. Bioresour Technol. 2018;264:359–369 https://doi.org/10.1016/J.BIORTECH.2018.06.013. Aryal et al., 2019 Aryal N, Wan L, Overgaard MH, et al. Increased carbon dioxide reduction to acetate in a microbial electrosynthesis reactor with a reduced graphene oxide-coated copper foam composite cathode. Bioelectrochemistry. 2019;128:83–93 https://doi.org/10.1016/j.bioelechem.2019.03.011. Batlle-Vilanova et al., 2015 Batlle-Vilanova P, Puig S, Gonzalez-Olmos R, Vilajeliu-Pons A, Balaguer MD, Colprim J. Deciphering the electron transfer mechanisms for biogas upgrading to biomethane within a mixed culture biocathode. RSC Adv. 2015;5:52243–52251 https://doi.org/10.1039/c5ra09039c. Batlle-Vilanova et al., 2019 Batlle-Vilanova P, Rovira-Alsina L, Puig S, et al. Biogas upgrading, CO2 valorisation and economic revaluation of bioelectrochemical systems through anodic chlorine production in the framework of wastewater treatment plants. Sci Total Environ 2019; https://doi.org/10.1016/j.scitotenv.2019.06.361.

Benjaminsson et al., 2013 Benjaminsson, G., Benjaminsson, J., Rudberg, R., 2013. 2Power-to-Gas – A technical review. Blasco-Gómez et al., 2017 Blasco-Gómez R, Batlle-Vilanova P, Villano M, Balaguer MD, Colprim J, Puig S. On the edge of research and technological application: a critical review of electromethanogenesis. Int J Mol Sci. 2017;18:1–32 https://doi.org/10.3390/ijms18040874. Bo et al., 2014 Bo T, Zhu X, Zhang L, et al. A new upgraded biogas production process: coupling microbial electrolysis cell and AD in single-chamber, barrelshape stainless steel reactor. Electrochem commun. 2014;45:67–70 https://doi.org/10.1016/j.elecom.2014.05.026. Buttler and Spliethoff, 2018 Buttler A, Spliethoff H. Current status of water electrolysis for energy storage, grid balancing and sector coupling via power-togas and power-to-liquids: a review. Renew Sustain Energy Rev. 2018;82:2440– 2454 https://doi.org/10.1016/j.rser.2017.09.003. Cerrillo et al., 2017 Cerrillo M, Vinas M, Bonmatí A. Startup of electromethanogenic microbial electrolysis cells with two different biomass inocula for biogas upgrading. ACS Sustain Chem Eng. 2017;5:8852–8859 https://doi.org/10.1021/acssuschemeng.7b01636. Cheng et al., 2009 Cheng S, Xing D, Call DF, Logan BE. Direct biological conversion of electrical current into methane by electromethanogenesis. Environ Sci Technol. 2009;43:3953–3958 https://doi.org/10.1021/es803531g. Christodoulou et al., 2017 Christodoulou X, Okoroafor T, Parry S, VelasquezOrta S. The use of carbon dioxide in microbial electrosynthesis: advancements, sustainability and economic feasibility. J CO2 Util. 2017;22:407–408 https://doi.org/10.1016/j.jcou.2017.11.006. Christodoulou and Velasquez-Orta, 2016 Christodoulou X, Velasquez-Orta SB. Microbial electrosynthesis and anaerobic fermentation: an economic evaluation for acetic acid production from CO2 and CO. Environ Sci Technol. 2016; acs.est.6b02101 https://doi.org/10.1021/acs.est.6b02101. Doud and Angenent, 2014 Doud DFR, Angenent LT. Toward electrosynthesis with uncoupled extracellular electron uptake and metabolic growth: enhancing current uptake with Rhodopseudomonas palustris. Environ Sci Technol Lett.

2014;1:351–355 https://doi.org/10.1021/ez500244n. Gahlot et al., 2020 Gahlot P, Ahmed B, Tiwari SB, et al. Conductive material engineered direct interspecies electron transfer (DIET) in AD: mechanism and application. Environ Technol Innov. 2020;20:101056 https://doi.org/10.1016/j.eti.2020.101056. Gajaraj et al., 2017 Gajaraj S, Huang Y, Zheng P, Hu Z. Methane production improvement and associated methanogenic assemblages in bioelectrochemically assisted AD. Biochem Eng J. 2017;117:105–112 https://doi.org/10.1016/j.bej.2016.11.003. Geppert et al., 2016 Geppert F, Liu D, van Eerten-Jansen M, Weidner E, Buisman C, ter Heijne A. Bioelectrochemical power-to-gas: state of the art and future perspectives. Trends Biotechnol 2016; https://doi.org/10.1016/j.tibtech.2016.08.010. Geppert et al., 2019 Geppert F, Liu D, Weidner E, Heijne A ter. Redox-flow battery design for a methane-producing bioelectrochemical system. Int J Hydrog Energy. 2019;44:21464–21469 https://doi.org/10.1016/j.ijhydene.2019.06.189. Hou et al., 2013 Hou J, Liu Z, Zhang P. A new method for fabrication of graphene/polyaniline nanocomplex modified microbial fuel cell anodes. J Power Sources. 2013;224:139–144 https://doi.org/10.1016/j.jpowsour.2012.09.091. Huskinson et al., 2014 Huskinson B, Marshak MP, Suh C, et al. A metal-free organic-inorganic aqueous flow battery. Nature. 2014;505:195–198 https://doi.org/10.1038/nature12909. IRENA, 2018 IRENA, 2018. Biogas for Road Vehicles Technology Brief. International Renewable Energy Agency, Abu Dhabi. Jin et al., 2017 Jin X, Zhang Y, Li X, Zhao N, Angelidaki I. Microbial electrolytic capture, separation and regeneration of CO2 for biogas upgrading. Environ Sci Technol. 2017;51:9371–9378 https://doi.org/10.1021/acs.est.7b01574. Jourdin et al., 2016a Jourdin L, Freguia S, Flexer V, Keller J. Bringing high-rate, CO2-based microbial electrosynthesis closer to practical implementation through improved electrode design and operating conditions. Environ Sci Technol.

2016a;50:1982–1989 https://doi.org/10.1021/acs.est.5b04431. Jourdin et al., 2016b Jourdin L, Lu Y, Flexer V, Keller J, Freguia S. Biologically induced hydrogen production drives high rate/high efficiency microbial electrosynthesis of acetate from carbon dioxide. ChemElectroChem. 2016b;3:581–591 https://doi.org/10.1002/celc.201500530. Kokkoli et al., 2018 Kokkoli A, Zhang Y, Angelidaki I. Microbial electrochemical separation of CO2 for biogas upgrading. Bioresour Technol. 2018;247:380–386 https://doi.org/10.1016/j.biortech.2017.09.097. Krieg et al., 2014 Krieg T, Sydow A, Schröder U, Schrader J, Holtmann D. Reactor concepts for bioelectrochemical syntheses and energy conversion. Trends Biotechnol. 2014;32:645–655 https://doi.org/10.1016/j.tibtech.2014.10.004. Kvist and Aryal, 2019 Kvist T, Aryal N. Methane loss from commercially operating biogas upgrading plants. Waste Manag. 2019;87:295–300 https://doi.org/10.1016/j.wasman.2019.02.023. Lee et al., 2019 Lee M, Nagendranatha Reddy C, Min B. in situ integration of microbial electrochemical systems into AD to improve methane fermentation at different substrate concentrations. Int J Hydrog Energy 2019;2380–2389 https://doi.org/10.1016/j.ijhydene.2018.08.051. Lin et al., 2015 Lin K, Chen Q, Gerhardt MR, et al. Alkaline quinone flow battery. Sci (80). 2015;349:1529–1532 https://doi.org/10.1126/science.aab3033. Liu et al., 2019 Liu C, Sun D, Zhao Z, Dang Y, Holmes DE. Methanothrix enhances biogas upgrading in microbial electrolysis cell via direct electron transfer. Bioresour Technol. 2019;291:121877 https://doi.org/10.1016/j.biortech.2019.121877. Liu et al., 2012 Liu F, Rotaru A-E, Shrestha PM, Malvankar NS, Nevin KP, Lovley DR. Promoting direct interspecies electron transfer with activated carbon. Energy Environ Sci. 2012;5:8982–8989 https://doi.org/10.1039/C2EE22459C. Liu et al., 2017 Liu SY, Charles W, Ho G, Cord-Ruwisch R, Cheng KY. Bioelectrochemical enhancement of AD: comparing single- and two-chamber

reactor configurations at thermophilic conditions. Bioresour Technol. 2017;245:1168–1175 https://doi.org/10.1016/j.biortech.2017.08.095. Morita et al., 2011 Morita M, Malvankar N, Franks A, et al. Potential for direct interspecies electron transfer in methanogenic. MBio. 2011;2:5–7 https://doi.org/10.1128/mBio.00159-11.Editor. Nevin et al., 2010 Nevin KP, Woodard TL, Franks AE, Summers ZM, Lovley DR. Microbial electrosynthesis: feeding microbes electricity to convert carbon dioxide and water to multicarbon extracellular organic compounds. MBio. 2010;1:e00103–e00110 https://doi.org/10.1128/mBio.00103-10. Park et al., 2018 Park J, Lee B, Tian D, Jun H. Bioelectrochemical enhancement of methane production from highly concentrated food waste in a combined anaerobic digester and microbial electrolysis cell. Bioresour Technol. 2018;247:226–233 https://doi.org/10.1016/j.biortech.2017.09.021. Reguera et al., 2005 Reguera G, McCarthy KD, Mehta T, Nicoll JS, Tuominen MT, Lovley DR. Extracellular electron transfer via microbial nanowires. Nature. 2005;435:1098–1101 https://doi.org/10.1038/nature03661. Rojas et al., 2018a Rojas MDPA, Mateos R, Sotres A, Zaiat M, Gonzalez ER, et al. Microbial electrosynthesis (MES) from CO2 is resilient to fluctuations in renewable energy supply. Energy Convers Manag. 2018a;177:272–279 https://doi.org/10.1016/j.enconman.2018.09.064. Rojas et al., 2018b Rojas MDPA, Zaiat M, Gonzalez ER, De Wever H, Pant D. Effect of the electric supply interruption on a microbial electrosynthesis system converting inorganic carbon into acetate. Bioresour Technol. 2018b;266:203– 210 https://doi.org/10.1016/j.biortech.2018.06.074. Rotaru et al., 2014 Rotaru AE, Shrestha PM, Liu F, et al. A new model for electron flow during AD: direct interspecies electron transfer to Methanosaeta for the reduction of carbon dioxide to methane. Energy Environ Sci. 2014;7:408–415 https://doi.org/10.1039/c3ee42189a. Schröder et al., 2015 Schröder U, Harnisch F, Angenent LT. Microbial electrochemistry and technology: terminology and classification. Energy Environ Sci. 2015;8:513–519 https://doi.org/10.1039/C4EE03359K.

Sun et al., 2015 Sun Q, Li H, Yan J, Liu L, Yu Z, Yu X. Selection of appropriate biogas upgrading technology—a review of biogas cleaning, upgrading and utilization. Renew Sustain Energy Rev. 2015;2015(51):521–532. Tremblay et al., 2017 Tremblay PL, Angenent LT, Zhang T. Extracellular electron uptake: among autotrophs and mediated by surfaces. Trends Biotechnol 2017; https://doi.org/10.1016/j.tibtech.2016.10.004. Xu et al., 2014 Xu H, Wang K, Holmes DE. Bioelectrochemical removal of carbon dioxide (CO2): an innovative method for biogas upgrading. Bioresour Technol. 2014;173:392–398 https://doi.org/10.1016/J.BIORTECH.2014.09.127. Yin et al., 2016 Yin Q, Zhu X, Zhan G, et al. Enhanced methane production in an AD and microbial electrolysis cell coupled system with co-cultivation of Geobacter and Methanosarcina. J Environ Sci (China). 2016;42:210–214 https://doi.org/10.1016/j.jes.2015.07.006. Zeppilli et al., 2017 Zeppilli M, Mattia A, Villano M, Majone M. Three-chamber bioelectrochemical system for biogas upgrading and nutrient recovery. Fuel Cell. 2017;17:593–600 https://doi.org/10.1002/fuce.201700048. Zeppilli et al., 2019 Zeppilli M, Simoni M, Paiano P, Majone M. Two-side cathode microbial electrolysis cell for nutrients recovery and biogas upgrading. Chem Eng J. 2019;466–476 https://doi.org/10.1016/j.cej.2019.03.119. Zhang et al., 2012 Zhang T, Nie H, Bain TS, et al. Improved cathode materials for microbial electrosynthesis. Energy Environ Sci. 2012;217–224 https://doi.org/10.1039/c2ee23350a.

Chapter 14

Photosynthetic biogas upgrading: an attractive biological technology for biogas upgrading

Vijay Kumar Garlapati¹, Swati Sharma¹ and Surajbhan Sevda², ¹1Department of Biotechnology and Bioinformatics, Jaypee University of Information Technology, Waknaghat, India, ²2Department of Biotechnology, National Institute of Technology Warangal, Warangal, India

Abstract

The future demands of transportation and industrial sectors necessitate groundbreaking research toward a sustainable energy source. Biogas from anaerobic digestion has been a well-studied research concept for the last two to three decades. The biogas technology suffers from the composition of unwanted contaminants (mainly CO2, H2S) in the production stream and utilization of cumbersome, energy-intensive removal technologies. The photosynthetic microalgae utilize the sequestered CO2 from biogas to synthesize different value-added products and aid biogas upgradation to attain the energy-rich biomethane. This chapter puts forth the positive attributes of photosynthetic microalgae-based biogas upgradation toward enhanced CO2 removal from biogas while meeting the natural grid system’s standards. Greater emphasis has been given toward photosynthetic biogas upgradation process set up, with possible integration with wastewater treatment and biomass production, the existing photobioreactors, influential process variables for better biogas upgradation, and discussion of the prospects of the photosynthetic biogas upgradation technology.

Keywords

Photosynthetic; microalgae; biogas upgradation; biomethane; photobioreactor

Chapter outline

Outline


14.2 Positive attributes of photosynthetic “microalgae” toward biogas upgradation 385

14.3 CO2 and H2S removal through photosynthetic-bacterial associated biogas upgradation 386

14.4 Microalgae-based biogas upgrading and concomitant wastewater treatment 387

14.5 Photobioreactor designs for biogas upgradation 388

14.6 Impact of different process variables in biogas upgradation 392

14.6.1 Light intensity 392 14.6.2 Media pH 394 14.6.3 Temperature 394 14.6.4 Biogas composition 395 14.6.5 Gas flow rate 395

14.7 The future prospects 396

14.8 Conclusion 401

References 401

14.1 Introduction

Biogas generated through anaerobic digestion serves as one of the energy sources for industry and society (Bose et al., 2020; Angelidaki et al., 2018). The projected anthropogenic-based atmospheric emissions of methane are expected to reach a mammoth 405 Tg/year by 2030 and will remain as a possible potent (>25 times) greenhouse gas (GHG) than CO2. Methane capture through the AD process seems to be a vital sustainable production method of methane through AD-based biogas upgradation technology [Nagarajan et al., 2019a; Angelidaki et al., 2018; environmental impact assessment (EIA)]. Biogas is the major product of AD with CH4 and CO2 fractions of 50%–70% and 30%–50%, respectively. In AD-based biogas, the compositions of CH4 and CO2 mainly depend on the feedstock and operational conditions, etc. The product stream of AD also includes different constituents such as nitrogen (N2, 0%–3%, due to the saturated air in the influent), water moisture (H2O, 5%–10%, at higher temperatures from medium evaporation), oxygen (O2, 0%–1%, from the substrate flow), hydrogen sulfide (H2S, 0–10,000 ppmv, by reduction of sulfate present in waste streams), ammonia (NH3, through proteinaceous materials or urine hydrolysis), hydrocarbons (0–200 mg/m³), and siloxanes (0–41 mg m³, from cosmetic medical effluents (Angelidaki et al., 2018; Muñoz et al., 2015). Generally, the calorific value of CH4 at sewage treatment plants is low, at around 50.4 MJ/kg or 36 MJ/m³, which is attributed to the presence of CO2 or N2 streams in the produced biogas (Mes et al., 2003). The extraordinarily corrosive and toxic H2S and NH3 can damage metal parts, heat, and power units through SO2 emissions (by combustion). The siloxanes presence in biogas damage biogas combustion engines and valves via sticky residues generation through silicone oxides combustion (Abatzoglou and Boivin, 2009). To enhance the calorific value and application range of biogas, removing pollutants is a prerequisite for a sustainable approach (Bose et al., 2020). Biogas upgrading differs from the “biogas cleaning” in the sense that the former aims to enhance the low calorific value of biogas (20–25 MJ/m³) toward the standards of natural gas (calorific value of near 50.76 MJ/m³) by removing the CO2 fraction, and the latter aims to remove the impurities/toxic compounds

(H2S, VOCs, siloxanes, Si, CO, and NH3) (Angelidaki et al., 2019; Kougias and Angelidaki, 2018; Aryal and Kvist, 2018). Biogas upgradation removes the CO2 and reduces the gas relative density, and enhances the Wobbe index (Meier et al., 2015; Ryckebosch et al., 2011). The generated biogas through the AD process can be utilized in combined heat and power (CHP) plants for heat and power generation through the desulfurizing step of crude biogas (Nagarajan et al., 2019a; Patterson et al., 2011). However, biogas’ methane content must be enhanced for a better calorific value that enables biogas use as a transportation fuel. Removing extraneous materials from crude biogas is needed to overcome the engine-related problems while using combustion engines (Nagarajan et al., 2019a). The upgraded biogas, that is, biomethane, can be easily integrated into natural gas grid injectors along with the transportation fuel with the added advantages of enhanced residual storage capabilities (through attained energy density) with higher efficiencies (Schmid et al., 2019). Based on the international legislation bodies, the composition of upgraded biogas, that is, biomethane, needs to consist of ≥95% CH4,≤2% CO2,≤0.3% O2, and H2S in negligible amounts (Toledo-Cervantes et al., 2016). For the utilization of upgraded biogas, that is, biomethane, in different natural gas grids, the CO2 levels need to be with in certain limits that differ with each country (up to 2.5% in , 3% in Austria, 4% in Sweden, and 6% in Switzerland and ) (Bose et al., 2020; Meier et al., 2015). The existing physicochemical-based biogas upgrading technologies for CO2 removal mainly work on the principles of adsorption (swing adsorption), membrane (membrane separation), and scrubbing (chemical/organic/water) principles, which need a prior H2S cleaning step (Rodero et al., 2019a). The technologies also suffer from different operational complexities, require high energy inputs (200–700 Wh/m³), and are associated with an acquisition cost of chemicals/organic substrates and digestion process, limiting the cost-effective replacement of natural gas with the upgraded biogas, that is, biomethane (Rodero et al., 2018). Recently, proposed cryogenic-based biogas upgradation yields pure methane and food-grade CO2 fractions but suffers from high investment costs (Song et al., 2018). Moreover, these technologies only separate the CO2 from CH4, but it is not accomplished with any CO2 converting mechanisms toward beneficial products and releases the separated CO2 into the environment (Schmid et al., 2019). The counterpart attempted biological approaches (biofiltration/in situ microaerobic AD, for H2S removal) with an integrated hydrogenotrophic step toward a biogas upgradation two-stage process and is only feasible in surplus sustainable electricity locations (Angelidaki et al.,

2018; Muñoz et al., 2015). Rigorous research attempts have been made toward sustainable biogas upgradation with the possible conversion of CO2 in to different valuable products and CO2 sequestration abilities and ended up with a solution called “photosynthetic-based biogas upgrading” (Bose et al., 2020). The photosynthetic machinery (microalgae) can sequester the available biogas CO2 and its further utilization for biomass production, which comprises a plethora of industrial commodities such as health, medical, pharmaceutical, food, and energy compounds (Bhatia et al., 2020; Gour et al., 2018, 2020; Sevda et al., 2019; Bajpai et al., 2017; Jha et al., 2017). This chapter puts forth the positive attributes of photosynthetic-based biogas upgradation technology and its simultaneous wastewater treatment (WWT) with concomitant biomass production and a discussion of the different process variables that affect biogas upgradation. Finally, the chapter ends with the possible cascading of photosynthetic biogas upgradation with the different algal biorefinery systems and discusses the technical challenges and prospects.

14.2 Positive attributes of photosynthetic “microalgae” toward biogas upgradation

The positive attributes of photosynthetic microalgae-based biogas upgradation include the ability of CO2 removal and its transformation toward biomass production with the help of captured solar energy. The positive attributes of microalgae include its faster growth rate and its adaptability to harsh process conditions (Moutinho et al., 2018). The biofuel products of algal biomass include biodiesel through produced lipids transesterification (Garlapati et al., 2021, 2017, 2013; Kumari et al., 2009), bioethanol (through fermentation of carbohydrates) (Sevda et al., 2019), and more biogas through AD of leftover algal biomass (Mohlin et al., 2018). Photosynthetic-based biogas upgradation also efficiently removes the CO2 (through CO2 sequestration) and H2S (through oxidation of H2S to SO/SO4²−, by synergistic association with sulfur-oxidizing bacteria) and is considered as a low-cost sustainable biogas upgradation approach (Rodero et al., 2019a). The overall advantage of microalgal-based biogas upgradation mainly lies in converting CO2 to energy and high-value industrial commodities using milder reaction conditions and subsequently developing a sustainable circular bioeconomy (Chandel et al., 2020; Sevda et al., 2020). The ideal characteristic of microalgae utilized for photosynthetic-based biogas upgradation includes its tolerance to methane and H2S in raw biogas to remove the CO2 (Nagarajan et al., 2019a). The tolerance can be achieved by developing mutant strains (Chlorella MM-2 and Chlorella sp. MB-9) through random mutagenesis, which can tolerate up to 80% methane and 100 ppm of H2S in the biogas stream (Kao et al., 2012a,b). The developed mutant strains enhance the biogas upgrading potential with moderate methane levels (Yan et al., 2014). The microalgal species of Chlorella, Spirulina, Nannochloropsis, Scenedesmus, and Chlorococcum are ideal species for photosynthetic-based biogas upgradation due to their rapid growth, and high tolerance to stress factors and CO2 (Asfouri et al., 2017).

14.3 CO2 and H2S removal through photosyntheticbacterial associated biogas upgradation

CO2 removal through photosynthetic-based biogas upgradation relies on fixating biogas CO2 by microalgae and using the fixed CO2 and solar radiation, obtaining the algal biomass. During the oxygenic photosynthesis (redox reaction), the electrons of water photolysis are used for the reduction of biogas CO2 [Eq. (14.1)] (Muñoz et al., 2015).

(14.1)

For oxygenic photosynthesis, mass transfer of CO2 from the biogas to the algal cultivation media has to occur. An appropriate design of photobioreactors (PBRs) and optimum process conditions is a prerequisite (Angelidaki et al., 2019). The presence of nutrients in cultivation media dictates the microalgal growth and plays a role in better sequestration results (Zhang et al., 2017). The byproduct (digests) of AD of organic waste, which comprise the N and P, serves as a cheap nutrient source for microalgae utilization for upgradation, thus reducing the operational costs of the process (Toledo-Cervantes et al., 2016). The high alkalinity of digestates favors CO2 mass transfer to the aqueous phase (Rodero et al., 2018). Chlorella sp. and Spirulina sp. (due to the tolerance to wide pH and CO2 concentrations) are the typically used microalgae in photosynthetic biogas upgradation technologies (Muñoz et al., 2015). The presence of H2S in the algal medium inhibits the microalgal growth by reducing the electron flow from photosystem I (PSI) to photosystem II (PSII), which eventually inhibits photosynthesis (Nagarajan et al., 2019a; GonzálezCamejo et al., 2017). The H2S removal in photosynthetic biogas upgradation is mediated by the sulfide-oxidizing bacteria (chemolithoautotrophs, synergically present in the microalgal-based biogas upgradation). The chemolithoautotrophic organisms obtain the energy from the oxidation of H2S or its byproducts to SO4²- with the help of electron acceptor (molecular oxygen/nitrate) [Eqs. (14.2) and (14.3), respectively] (Franco-Morgado et al., 2017). Under the limitations of the electron acceptor, the elemental sulfur is formed [Eqs. (14.2) and (14.4), respectively] (López et al., 2018).

(14.2)

Aerobic oxidation

(14.3)

(14.4)

The byproducts of H2S oxidation in microalgal-based biogas upgradation, that is, elemental sulfur (S) and sulfate (SO4²-), have nonhazardous natures. The elemental sulfur can be separated through the physical method and valorized further (Franco-Morgado et al., 2017), whereas microalgae can utilize sulfate for growth purposes.

14.4 Microalgae-based biogas upgrading and concomitant wastewater treatment

Photosynthetic-based biogas upgradation can be easily integrated with WWT, and algal biomass production as the photosynthetic microalgae can remove the nitrogen and phosphorus wastewater streams and be utilized for algal growth. The removed chemical oxygen demand (COD) and BOD from wastewater streams coupled with fixed CO2 can be used to grow greater algal biomass, which consists of energy-rich and value-added products (Mandova et al., 2018). The photosynthetic-based biogas upgradation through microalgae-bacterial association through the symbiotic hierarchy of photosynthetic microalgae and heterotrophic bacteria stands alone as an ideal biotechnological approach for simultaneous removal of biogas pollutants such as CO2 and H2S (Rodero et al., 2019a; Bahr et al., 2014). This novel photosynthetic biogas upgrading technology relies on the heuristics of CO2 fixation by solar-driven photosynthesis coupled with bacterial-based oxidation of H2S to SO4²- with the utilization of photosynthetically produced O2 (Angelidaki et al., 2019). An ideal photosynthetic-based biogas upgradation integrated WWT and biomass production comprising a two-step process with a separate biomass harvesting step is shown in Fig. 14.1. The first step of the integrated process aims to upgrade biogas by removing the CO2 that meets natural grid standards. The second step takes care of CO2 sequestration through the fixation of captured CO2 by microalgae. The final step is CO2 utilization for microalgal biomass production and its harvesting for utilization in biochemicals and biofuels production (Fig. 14.1). The photosynthetic-based biogas upgradation can obtain better results with the selection of better microalgae species with specific properties and the optimization of system parameters (Bose et al., 2019). The recent research findings related to the microalgal-based biogas upgradation and concomitant WWT are summarized in Table 14.1.

Figure 14.1 Process flow diagram of an ideal photosynthetic-based biogas upgradation system integrated with wastewater treatment with value-added biomass production facilities. Modified from Bose, A., Lin, R., Rajendran, K., O’Shea, R., Xia, A., Murphy, J.D., 2019. How to optimise photosynthetic biogas upgrading: a perspective on system design and microalgae selection. Biotechnol. Adv. 37 (8), 107444.

Table 14.1

Biogas composition

Microalgae/microalgae–coculti

–

Chlorella vulgaris cultivated w

Raw biogas (CH4—61.23%, CO2—34.69%)

C. vulgaris cocultivated with a

Synthetic biogas (70%CH4, 29.5% CO2, 0.5% H2S)

Microalgae and bacteria

Crude biogas (CH4—64.59% CO2—33.79% H2S—<0.01%)

C. vulgaris cocultivated with G

Synthetic biogas (69.5%CH4, 30% CO2, 0.5% H2S)

Picochlorum sp. cultivated wi

Crude biogas (CH4—64.59% CO2–33.79% H2S—<0.01%)

C. vulagris with activated slud

Synthetic biogas with various concentrations of carbon dioxide


Synthetic biogas (70%CH4, 29.5% CO2, 0.5% H2S)

Chlorella minutissima

Synthetic biogas (CH4—62.38% CO2—31.19%)





C. vulgaris cocultivated with a

Raw biogas (CH4—57.32% CO2—39.25%)

C. vulgaris cocultivated with f

Crude biogas (CH4—63.84% CO2—31.02%)

Chlorella sp.

Raw biogas (CH4—70.7% CO2—26.1% H2S—1550ppm)

Scendesmus sp.

Synthetic raw biogas (CH4–60.55% CO2—38.02%)

Chlorella sp.


C. vulgaris


Nannochloropsis Oleoabunda

Raw biogas (CH4—61.75% CO2—35.28%)

Scendesmus obliquus

Synthetic biogas (70% CH4, 29.5% CO2, 0.5% H2S)

C. vulgaris


Scendesmus sp.

Raw biogas (CH4—58.67% CO2—37.54%)

Scendesmus obliquus

Raw biogas (CH4—67.35%, CO2—28.41% H2S—<0.005%) of 96 L total volume

Chlorella sp.

14.5 Photobioreactor designs for biogas upgradation

The most common configurations of bioreactors used for photosynthetic biogas upgradation include high-rate algal ponds (HRAPs) and closed PBR. Among these, the capital investment and operational costs associated with HRAPs are less than those of closed bioreactors (Torres et al., 2020). The HRAPs suffer from the drawbacks of low biomass concentrations and productivities (due to insufficient photosynthetic activity), and large land requirements with high water footprints (Posadas et al., 2017). In contrast, the closed photobioreactors systems can attain biomass concentrations of 2–8 g/TSS/L and 25–45 g/m² productivities (i.e., 5–6 times higher than HRAPs) on continuous mode due to the presence of enhanced light utilization efficiencies (4–6%), high turbulent conditions, and more illuminated surface-to-volume ratio over HRAPs (Bose et al., 2019, 2020; Vo et al., 2019). The common PBRs utilized for photosynthetic biogas upgrading include plate, air-lift (bubble column), and tubular PBRs, which are depicted in Fig. 14.2.

Figure 14.2 Simplistic schematic representation of closed photobioreactors in continuous operation mode (A) plate-type; (B) air lift (bubble column) algal bioreactor; (C) tubular type algal photobioreactor.

The flat-plate PBRs are robust, high surface-to-volume ratio configurations with easy operation and superior efficiency. The low reactor thickness of plate-type PBRs helps attain greater yields (5–20 times more) than other closed PBR counterparts (Vo et al., 2019). A novel plate PBR, namely thin-film flat-plate PBR, was proposed to reduce the cost and improve scalability (Yan and Zheng, 2013). The drawbacks associated with plate-type PBR scalability include high cost of constructive materials, enhanced system costs related to the optimal plate spacing for minimizing the shading of reactors, and greater space requirements (Bose et al., 2019). The positive attributes of airlift (bubble column, made from glass) PBRs include effective temperature control and proper design, efficient mixing and mass transfer, good pH control, and low power requirements. The airlift PBRs suffer from the requirement of optimized column diameter for light control, attainment of moderate surface-to-volume ratios, high capital costs, and scalability issues (Endres et al., 2018; Guo et al., 2017a,b; Huang et al., 2017b). The tubular PBRs used for photosynthetic biogas upgradation suffer from problems in temperature tuning during winters. The large-diameter tubes in the configuration cause problems in controlling the light, adequate mixing and mass transfer, high power requirement, and a less effective pH control system due to poor mixing. The positive attributes of tubular PBRs include their robustness, high surface-to-volume ratio, and limited scalability applications (compared with bubble column PBRs) (Nag Dasgupta et al., 2010; Sierra et al., 2008; Vasumathi et al., 2012). Another major constraint of closed PBR is oxygen concentration build-up through photosynthesis, which imparts a high oxygen content to the upgraded biogas (Torres et al., 2020). The higher oxygen constraint can be overcome by directly introducing the biogas to the PBRs or utilizing the external absorption column where the CO2-containing algal media have to recirculate to the PBR (Meier et al., 2015). Chlorella sp., Pseudanabaena sp., and Scenedesmus sp. are the most commonly employed microalgal species. Thiobacillus sp. is the widely utilized bacterial species mostly found in algal-bacterial PBRs with the goal of photosynthetic biogas upgrading (Angelidaki et al., 2019; Toledo-Cervantes et

al., 2017a,b).

14.6 Impact of different process variables in biogas upgradation

Algal growth plays a predominant role in photosynthetic-based biogas upgradation, and it mainly depends on the different process variables of the upgradation process. The process variables, namely, temperature, light wavelength and intensity, gas flow rate, and biogas composition, play an important role in achieving enhanced upgrading results (Angelidaki et al., 2018; Muñoz and Guieysse, 2006).

14.6.1 Light intensity

Light provides an energy source for the microalgal photosynthesis related to microalgal growth and is a constituent of the primary metabolism of microalgae. Light intensity is also a critical factor in regulating microalgal lipids and valueadded product production (Srinuanpan et al., 2020; Zhao et al., 2013). Hence, during low-light intensities, increased photoperiods are recommended, which helps in attaining the microalgal growth, but high light intensity (above a certain threshold) also causes a photoinhibition effect (destroying the PS I and II) (Yan et al., 2016a; Jeong et al., 2013). The relationship between the specific growth rate (µ) and photosynthetic photon flux density (PPFD) (I) is shown in Eq. (14.5).

(14.5)

where µmax denotes the maximum specific growth rate and KE is the light halfsaturation constant for light (Lee et al., 2015; Kurano and Miyachi, 2005). Studies into the effect of light intensity and photoperiod on microalgal biogas upgradation have suggested that moderate light intensities of 150– 350 mmol/m²/s and photoperiods of 12–14 h help achieve enhanced CO2 removal from algal-based biogas upgrading systems (Yan et al., 2016a; Ouyang et al., 2015). Moreover, the light intensity and photoperiods vary with the microalgal cell density utilized in the photosynthetic biogas upgradation systems. Hence, higher light intensity is recommendable to tackle the light limitation phenomenon, and a steep increase in light intensity based on the growth curve of the microalgae will give positive results in microalgal-based biogas upgradation processes for CO2 removal (Srinuanpan et al., 2018; Das et al., 2011; Pilon et al., 2011; Yan et al., 2013b). Several researchers have reported that high light intensity helps in achieving maximum cell growth but reduces the algal biomass content (lipids) due to utilization of light energy it for microalgal cell division rather than utilizing it for biomass accumulation (George et al., 2014; Cheirsilp and Torpee, 2012; Yeesang and Cheirsilp, 2011). An enhanced CO2 removal was reported with the utilization of red-light wavelength (600–700 nm) in photosynthetic-based biogas upgradation systems due to the better absorption of red-light wavelength by chlorophylls in microalgal cells (Yan et al., 2016b; Kim et al., 2013; Ho et al., 2011). Zhao et al. (2013) and Yan et al. (2016b) suggested using red-light wavelength with other light wavelengths to avoid the light saturation and photoinhibition effects in photosynthetic-based biogas upgradation systems. The flexibility of regulating the light wavelengths associated with artificial light (LED) sources was reported as promising in photosynthetic biogas upgradation systems rather than natural solar light (Yan et al., 2016b). Promising results with moderate light intensities in photosynthetic biogas upgradation systems were reported with Leptolyngbya sp. CChF1 (Nagarajan et al., 2019a; Choix et al., 2017) and Scenedesmus sp. (Ouyang et al., 2015), which results in enhanced methane content and concomitant enhanced nutrient (COD, TN, and TP) removal from the slurry. Yan and Zheng (2013) and Yan et al. (2016a) reported promising CO2 removal strategies with Chlorella sp. by utilizing moderate light intensity (350 μmol/m²/s with a 14 h:10 h light/dark period, 86.15% CO2 removal) and low light intensity (300 μmol/m²/s, 16 h:8 h light/dark period, 81.6% CO2 removal). The utilization of red and blue light wavelengths (5:5) also gave fruitful results with the photosynthetic biogas upgradation systems toward enhanced CO2 removal (Yan

et al., 2016b; Zhang et al., 2017; Yan and Zheng, 2014; Yan et al., 2014).

14.6.2 Media pH

The culture media pH is one of the detrimental factors for the growth and metabolism of microalgae. The optimum pH varies with the algal strain, and can interfere with the nutrient uptake by microalgae. The media pH dictates the available form of inorganic carbon as bicarbonates or CO2 (Nagarajan et al., 2019b; Juneja et al., 2013; Markou et al., 2014). As the medium pH is alkaline, the inorganic carbon is available as bicarbonate that can be effectively utilized by microalgae. Generally, photosynthetic biogas upgrading favors an alkaline pH, enhancing the solubility of CO2 from biogas and remaining as dissolved carbonate in the medium that gives pure biogas (Nolla-Ardèvol et al., 2015). The medium pH has to be maintained as alkaline to better help of the microalgal growth. The volatile fatty acids (VFAs) taken by microalgae can drastically reduce the pH, inhibiting microbial growth (Chen et al., 2018). An alkaline pH microalgal system helps overcome the extraneous contamination problem and can trigger microalgae’s stress conditions toward enhanced lipid accumulation by microalgae (Tan et al., 2016; Bartley et al., 2014). The alkaline pH conditions of algal media contribution toward enhanced biogas upgradation and nutrient removal (WWT) has been reported by several researchers while utilizing Chlorella sp. (Cho et al., 2015; Lebrero et al., 2016; Ayre et al., 2017; Tongprawhan et al., 2014), Spirulina sp. (Nolla-Ardèvol et al., 2015), and Arthrospira platensis (Converti et al., 2009). Alkaline pH media favoring enhanced microalgal biogas upgradation has been reported with the alkaliphilic microalgal-bacterial consortium (Picochlorum sp., Thiobacillus sp., and Halospirulina sp.), where 91.5% CO2 removal efficiency was attained with an alkaline media pH 9.3 (Franco-Morgado et al., 2017). The utilization of microalgal consortia with alkaline pH media also yields enhanced CO2 removal efficiencies with a tubular photobioreactor (91.9% at pH 9.3) (Marín et al., 2019) and algal pond (97.8% at pH 9.7) (Rodero et al., 2018). The positive results with alkaline pH media are suggested to implement different photosynthetic biogas upgradation for enhanced CO2 removal.

14.6.3 Temperature

The temperature used in the photosynthetic biogas upgrading systems is mainly chosen based on the optimal temperature where the microalgae grow as the biogas from the AD process runs at ambient temperatures (25°C–30°C) only. The temperature of the biogas upgradation process influences only the biomass growth. It does not play an essential role in the biogas upgradation process, which is evident from the algal biogas upgrading with Leptolyngbya sp. (Choix et al., 2017). The results from a Chlorella sorokiniana-based biogas upgradation study suggested the inverse relation of algal system temperature with biogas CO2 solubility (Meier et al., 2017; Turon et al., 2016).

14.6.4 Biogas composition

Biogas composition (variations of CO2, CH4, and H2S) affects microalgae cultivation for photosynthetic biogas upgradation (Srinuanpan et al., 2020; González-Sánchez and Posten, 2017). The research results have suggested that the CO2 and H2S tolerance limits of microalgae are species-dependent (Wang et al., 2014; Amaro et al., 2011). The suitable microalgal strains for photosynthetic biogas upgradation studies include the CO2-tolerant Chlorella sp. and Scenedesmus sp. (Srinuanpan et al., 2017; Tongprawhan et al., 2014). The high concentration of CO2 results in low media pH, which inhibits the microalgal growth, and affects the CO2 sequestration abilities (Tongprawhan et al., 2014). Eq. (14.6) shows the relationship between the algal growth rate and the initial dissolved inorganic carbon, CO2, which looks like the Monod model equation (Wang et al., 2017a,b; Yan et al., 2014).

(14.6)

where µ denotes the specific growth rate, µmax is the maximum specific growth rate, Ks is the half-saturation constant, and “S” is the CO2 concentration. The biogas CH4 content dictates the microalgal growth in the biogas upgradation systems and CO2 sequestration abilities of the microalgae also (Srinuanpan et al., 2020). The CH4 tolerance limits vary with microalgae; Nannochloropsis gaditana tolerates a CH4 content of 50%–100% (Meier et al., 2015), whereas Chlorella sp. MM-2 (Kao et al., 2012a) and Chlorella sp. (Yan et al., 2014) tolerate around 20%–50% CH4 and 45%–55% CH4 contents, respectively. Another comparative study by Srinuanpan et al. (2017) found that marine Chlorella sp. and Scenedesmus sp. tolerate 40% CO2 in the biogas. The H2S content in biogas streams ranged from 0.005% to 2% (Srinuanpan et al., 2020). The research results on the effect of H2S on microalgae performance suggest that Chlorella sp. tolerate H2S levels up to 150–200 ppm (GonzálezSánchez and Posten, 2017; Kao et al., 2012b). Some algal species can convert H2S to sulfate, which was assimilated as the growth-limiting substrate under low pH conditions (Mera et al., 2016).

14.6.5 Gas flow rate

The gas flow rate plays a vital role in photosynthetic biogas upgradation processes related to possible CO2 removal with concomitant biomass production (Srinuanpan et al., 2020). The optimal flow rate helps in proper mixing and adequate CO2 supply for possible microalgal growth (Yan et al., 2016a). The enhanced gas flow rates help in attaining maximal biomass productivity and CO2 removal efficiencies, but with gas flow rates higher than the threshold limits of microalgal consumption, the culture pH becomes too acidic, which profoundly affects the nutrient availability, algal metabolism, CO2 transport, and especially, the photosynthesis (Tang et al., 2011).

The excess gas flow rates increase the size of gas bubbles, which reduces the specific surface area per gas volume and increases the culture medium (Srinuanpan et al., 2018; Su et al., 2017). The step-wise increment of biogas flow rates to photosynthetic biogas upgradation systems is the ideal approach for overcoming the harmful effect of excess gas flow rates, which facilitates the lower gas flow rates to the low algal cell density (initial conditions), and the flow rate will be increased with increasing cell density (with subsequent algal growth). The step-wise biogas flow rate also gives fruitful results with photosynthetic biogas upgradation systems in of CO2 removal with concomitant algal lipid production (Widjaja et al., 2009; Binnal and Babu, 2017). The biogas upgradation study of a 180 L photobioreactor using algalbacterial symbiosis revealed that an L/G ratio of 10 (among tested 0.5
Table 14.2

Algal reactor configurations

16.8 L glass photobioreactor filled with slurry and addition of different concentrations of strigolactone GR24 Biogas com Tubular photobioreactor Synthetic biogas composition 70% CH4, 29.5% CO2, 0.5% H2S.

Open 50 L photobioreactor Connected with bubble column operated continuous lyInjected biogas—65±1.5 % CH4; 32.0

HRAP attached with bubble column via recirculation broth Biogas mixture contains—70% CH4, 29.5% CO2, 0.5% H2S

Outdoor high rated algal pond attached with bubble absorption column Biogas composition—CO2 29.5%, H2S 0.5% an 25 L high rated algal pond connected with absorption bubble column Synthetic biogas 30% CO2, 0.5% H2S Two interconnected cylindrical glass incorporated with biogas and slurry with 60–70% methane content Two connected cylindrical glass have synthetic biogas with varying CO2 (25.27%–55.17%) concentration and slurry

Closed glass Photo bioreactor of 30 cm height and 20 cm diameter used Biogas composition—65%–73% CH4, 20%–25 Photo bioreactor with 20 L crude gas and 4 L slurry H2S— <50ppm 75 L photobioreactor connect with bubble column filled with real biogas 72% CH4, 28% CO2 180 L HRAP connected with bubble column Synthetic biogas— 30% CO2, 69.5% CH4, 0.5% H2S

Interconnected 180 L raceway photobioreactor with bubble column filled with synthetic biogas—30% CO2, 69.5% N2 a Raceway plant with recirculation rate 0.22 m/s filled with 10.6% CO2 flue gas Bubble column photobioreactor Biogas composition—69% CH4, 20% CO2 and 0.005% H2S 420 L raceway connect with bubble column reactor Bubble column reactor filled with real biogas 38%–80% CH4, 19%–62% CO2, and 0.2% H2S Tubular photobioreactor Biogas composition—41% CO2, 57.5% CH4 and 0.05% H2S 15 L Algal pond with absorption unit filled with real biogas 44%–48% CO2,55%–71% CH4 and 1% H2S Column photobioreactor incorporated with real biogas has compositions—70%–72% CH4, 17%–19% CO2

Raceway outdoor pilot plant with artificial biogas Consists of 40% CO2, 60% nitrogen

14.7 The future prospects

The primary constraints in photosynthetic biogas upgradation lie in commercialization of the technology, which involves a myriad of problems related to microalgal technology (such as developing CO2- and H2S-tolerant microalgal strains and cultivation problems), process engineering tasks, and integration with other value chains toward cost-effectiveness (Sun et al., 2016; Wang et al., 2016). The effectiveness in photosynthetic biogas upgradation lies in employing the CO2-, H2S-, and CO-tolerant microalgal species to better sequestration/removal of CO2 (Bose et al., 2020; Nagarajan et al., 2019a). More emphasis has to be made toward developing a potent microalgal strain, which can work with different streams of flue gases and hazardous industrial effluents (Sharma and Garlapati, 2021) and using genetic engineering and synthetic biology concepts. The AD slurries with high COD levels have a profound effect on the penetration of light in algal cultures, which can be overcome by utilizing the prior pretreatment of AD slurries (for removal of suspended solids and particulate matter), practicing of hetero-/mixo-tropic cultivations (without light energy), and using diluted slurries. The problem of having toxic and unwanted VFAs in the slurry can be reduced by adopting light energy, which reduces VFA uptake by microalgae. The energy and cost incurred by sterilization of AD slurries can be tackled by utilizing the alkaline cultivation media and selecting a robust algal strain which out-competes the competent bacteria out of the system (Nagarajan et al., 2019a). The advancements in process engineering concepts of photosynthetic biogas upgrading mainly lie in deg cost-effective, energyefficient PBRs with more flexible light sources, CO2 transfer, mixing, and cleaning mechanisms (Sevda et al., 2021). The integration of systems with novel microfluidic technologies with more sustainable green solvent-based downstream processing helps attain pure methane that can be easily integrated with natural gas grid systems (Bose et al., 2020; Samudrala et al., 2017). The implementation of microfluidics in algal cultivation and value-added products synthesis helps integrate algal-based biogas upgradation technologies with other bioelectrochemical systems toward cost-effectiveness and sustainable proof-ofconcept technology (Banerjee et al., 2019; Sharma et al., 2019; Sevda et al., 2019). The sustainability of photosynthetic biogas upgradation needs detailed

life cycle analysis studies to show the positive impact on climate change and to return the older AD-based biogas production systems to greater heights (Bose et al., 2019, 2020; Garlapati et al., 2019).

14.8 Conclusion

Photosynthetic biogas upgradation is a sustainable approach for removing CO2 for better calorific methane with possible integration of biomass growth toward value-added products. The technology operates under mild conditions with flexible monitoring conditions associated with PBRs. The photosynthetic microalgae with CO2 sequestration abilities can help in transforming the produced biogas (through AD) to energy-rich upgraded methane through the environmental-friendly approach. The technology requires in-depth research toward developing potent microalgae and proof-of-concept studies and commercialization toward biogas utilization in natural grids and transportation sector domains. Overall, this chapter summarizes the positive attributes, mechanisms for CO2 and H2S removal, existing PBRs, integration strategies with WWT, and value-added product streams, detrimental process variables, and technical challenges coupled with prospects for photosynthetic-based biogas upgradation.

References

Abatzoglou and Boivin, 2009 Abatzoglou N, Boivin S. A review of biogas purification processes. Biofuels, Bioprod Bioref. 2009;3(1):42–71. Amaro et al., 2011 Amaro HM, Guedes AC, Malcata FX. Advances and perspectives in using microalgae to produce biodiesel. Appl Energy. 2011;88(10):3402–3410. Ángeles et al., 2020 Ángeles R, Arnaiz E, Gutiérrez J, et al. Optimization of photosynthetic biogas upgrading in closed photobioreactors combined with algal biomass production. J Water Process Eng. 2020;38:101554. Angelidaki et al., 2018 Angelidaki I, Treu L, Tsapekos P, et al. Biogas upgrading and utilization: current status and perspectives. Biotechnol Adv. 2018;36(2):452–466. Angelidaki et al., 2019 Angelidaki I, Xie L, Luo G, et al. Biogas Upgrading: current and emerging technologies. In: Pandey A, Larroche C, Dussap CG, Gnansounou E, Khanal SK, Ricke S, eds. Biomass, Biofuels, Biochemicals, Biofuels: Alternative Feedstocks and Conversion Processes for the Production of Liquid and Gaseous Biofuels. Academic Press 2019;817–843. Aryal and Kvist, 2018 Aryal N, Kvist T. Alternative of biogas injection into the Danish gas grid system—a study from demand perspective. ChemEngineering. 2018;2:43. Asfouri et al., 2017 Asfouri N, Djalt H, Maroc F, Lamara S, Baba H, Abi-Aya S. Cultivation of marine microalga Nannochloropsis gaditana under various temperatures and nitrogen treatments: effect on growth, lipid and pigment content. Int J Biosci. 2017;10(3):209–216. Ayre et al., 2017 Ayre JM, Moheimani NR, Borowitzka MA. Growth of microalgae on undiluted anaerobic digestate of piggery effluent with high ammonium concentrations. Algal Res. 2017;24:218–226.

Bahr et al., 2014 Bahr M, Díaz I, Dominguez A, Gonzalez Sanchez A, Muñoz R. Microalgal-biotechnology as a platform for an integral biogas upgrading and nutrient removal from anaerobic effluents. Environ Sci Technol. 2014;48(1):573–581. Bajpai et al., 2017 Bajpai A, Garlapati VK, Gour RS, Kant A. Evaluation of microalgae from himalayan region for nutraceutical activities. Int J Pharma Bio Sci. 2017;8(2):174–178. Banerjee et al., 2019 Banerjee R, Kumar SJ, Mehendale N, Sevda S, Garlapati VK. Intervention of microfluidics in biofuel and bioenergy sectors: technological considerations and future prospects. Renew Sustain Energy Rev. 2019;101:548– 558. Bartley et al., 2014 Bartley ML, Boeing WJ, Dungan BN, Holguin FO, Schaub T. pH effects on growth and lipid accumulation of the biofuel microalgae Nannochloropsis salina and invading organisms. J Appl Phycol. 2014;26(3):1431–1437. Bhatia et al., 2020 Bhatia, L., Bachheti, R.K., Garlapati, V.K., Chandel, A.K., 2020. Third-generation biorefineries: a sustainable platform for food, clean energy, and nutraceuticals production. Biomass Convers. Biorefinery, pp. 1–16. Binnal and Babu, 2017 Binnal P, Babu PN. Statistical optimization of parameters affecting lipid productivity of microalga Chlorella protothecoides cultivated in photobioreactor under nitrogen starvation. South Afr J Chem Eng. 2017;23:26– 37. Bose et al., 2019 Bose A, Lin R, Rajendran K, O’Shea R, Xia A, Murphy JD. How to optimise photosynthetic biogas upgrading: a perspective on system design and microalgae selection. Biotechnol Adv. 2019;37(8):107444. Bose et al., 2020 Bose A, O’Shea R, Lin R, Murphy JD. A perspective on novel cascading algal biomethane biorefinery systems. Bioresour Technol. 2020;304:123027. Cao et al., 2017 Cao W, Wang X, Sun S, Hu C, Zhao Y. Simultaneously upgrading biogas and purifying biogas slurry using cocultivation of Chlorella vulgaris and three different fungi under various mixed light wavelength and photoperiods. Bioresour Technol. 2017;241:701–709.

Chandel et al., 2020 Chandel AK, Garlapati VK, Jeevan Kumar SP, Hans M, Singh AK, Kumar S. The role of renewable chemicals and biofuels in building a bioeconomy. Biofuels Bioprod Bioref. 2020;14:830–844. Cheirsilp and Torpee, 2012 Cheirsilp B, Torpee S. Enhanced growth and lipid production of microalgae under mixotrophic culture condition: effect of light intensity, glucose concentration and fed-batch cultivation. Bioresour Technol. 2012;110:510–516. Chen et al., 2018 Chen Y d, Ho SH, Nagarajan D, Ren N q, Chang JS. Waste biorefineries—integrating anaerobic digestion and microalgae cultivation for bioenergy production. Curr Opin Biotechnol. 2018;50:101–110. Cho et al., 2015 Cho HU, Kim YM, Choi YN, Xu X, Shin DY, Park JM. Effects of pH control and concentration on microbial oil production from Chlorella vulgaris cultivated in the effluent of a low-cost organic waste fermentation system producing volatile fatty acids. Bioresour Technol. 2015;184:245–250. Choix et al., 2017 Choix FJ, Snell-Castro R, Arreola-Vargas J, Carbajal-López A, Méndez-Acosta HO. CO2 removal from biogas by cyanobacterium Leptolyngbya sp. CChF1 isolated from the Lake Chapala, Mexico: optimization of the temperature and light intensity. Appl Biochem Biotechnol. 2017;183(4):1304–1322. Conde et al., 1993 Conde JL, Moro LE, Travieso L, et al. Biogas purification process using intensive microalgae cultures. Biotechnol Lett. 1993;15(3):317– 320. Converti et al., 2009 Converti A, Oliveira RPS, Torres BR, Lodi A, Zilli M. Biogas production and valorization by means of a two-step biological process. Bioresour Technol. 2009;100(23):5771–5776. Das et al., 2011 Das P, Lei W, Aziz SS, Obbard JP. Enhanced algae growth in both phototrophic and mixotrophic culture under blue light. Bioresour Technol. 2011;102(4):3883–3887. De Godos et al., 2014 De Godos I, Mendoza JL, Acién FG, et al. Evaluation of carbon dioxide mass transfer in raceway reactors for microalgae culture using flue gases. Bioresour Technol. 2014;153:307–314.

del Rosario Rodero et al., 2018 del Rosario Rodero M, Posadas E, ToledoCervantes A, Lebrero R, Muñoz R. Influence of alkalinity and temperature on photosynthetic biogas upgrading efficiency in high rate algal ponds. Algal Res. 2018;33:284–290. Doušková et al., 2010 Doušková I, Kaštánek F, Maléterová Y, Kaštánek P, Doucha J, Zachleder V. Utilization of distillery stillage for energy generation and concurrent production of valuable microalgal biomass in the sequence: biogascogeneration-microalgae-products. Energy Convers Manage. 2010;51(3):606– 611. Endres et al., 2018 Endres CH, Roth A, Bru TB. Modeling microalgae productivity in industrialscale vertical flat photobioreactors. Environ Sci Technol. 2018;52(9):5490–5498. Franco-Morgado et al., 2017 Franco-Morgado M, Alcántara C, Noyola A, Muñoz R, González-Sánchez A. A study of photosynthetic biogas upgrading based on a high rate algal pond under alkaline conditions: influence of the illumination regime. Sci Total Environ. 2017;592:419–425. Franco-Morgado et al., 2018 Franco-Morgado M, Toledo-Cervantes A, González-Sánchez A, Lebrero R, Muñoz R. Integral (VOCs, CO2, mercaptans and H2S) photosynthetic biogas upgrading using innovative biogas and digestate supply strategies. Chem Eng J. 2018;354:363–369. Gao et al., 2018 Gao S, Hu C, Sun S, Xu J, Zhao Y, Zhang H. Performance of piggery wastewater treatment and biogas upgrading by three microalgal cultivation technologies under different initial COD concentration. Energy. 2018;165:360–369. Garlapati et al., 2013 Garlapati VK, Kant R, Kumari A, Mahapatra P, Das P, Banerjee R. Lipase mediated transesterification of Simarouba glauca oil: a new feedstock for biodiesel production. Sustain Chem Process. 2013;1(1):11. Garlapati et al., 2017 Garlapati VK, Gour RS, Sharma V, et al. Current status of biodiesel production from microalgae in India. In: Austin: Wiley –Scrivener Publishing House; 2017;129–154. Singh L, Chaudhary G, eds. Advances in Biofeedstocks and Biofuels: Production Technologies for Biofuels. 2. Garlapati et al., 2019 Garlapati VK, Tewari S, Ganguly R. LCA of first-, second-

generation, and microalgae biofuels. In: Hosseini M, ed. Advances in Feedstock Conversion Technologies for Alternative Fuels and Bioproducts. Philadelphia, PA: Elsevier, Imprint: Woodhead Publishing; 2019;355–371. Garlapati et al., 2021 Garlapati VK, Mohapatra SB, Mohanty RC, Das P. Transesterified Olax scandens oil as a bio-additive: production and engine performance studies. Tribol Int. 2021;153:106653. George et al., 2014 George B, Pancha I, Desai C, et al. Effects of different media composition, light intensity and photoperiod on morphology and physiology of freshwater microalgae Ankistrodesmus falcatus–a potential strain for bio-fuel production. Bioresour Technol. 2014;171:367–374. González-Camejo et al., 2017 González-Camejo J, Serna-García R, Viruela A, et al. Short and long-term experiments on the effect of sulphide on microalgae cultivation in tertiary sewage treatment. Bioresour Technol. 2017;244:15–22. González-Sánchez and Posten, 2017 González-Sánchez A, Posten C. Fate of H2S during the cultivation of Chlorella sp deployed for biogas upgrading. J Environ Manage. 2017;191:252–257. Gour et al., 2018 Gour RS, Bairagi M, Garlapati VK, Kant A. Enhanced microalgal lipid production with media engineering of potassium nitrate as a nitrogen source. Bioengineered. 2018;9(1):98–107. Gour et al., 2020 Gour RS, Garlapati VK, Kant A. Effect of salinity stress on lipid accumulation in Scenedesmus sp and Chlorella sp.: feasibility of stepwise culturing. Curr Microbiol. 2020;77(5):779–785. Guo et al., 2017a Guo G, Cao W, Sun S, Zhao Y, Hu C. Nutrient removal and biogas upgrading by integrating fungal–microalgal cultivation with anaerobically digested swine wastewater treatment. J Appl Phycol. 2017a;29(6):2857–2866. Guo et al., 2017b Guo Y, Yuan Z, Xu J, et al. Metabolic acclimation mechanism in microalgae developed for CO2 capture from industrial flue gas. Algal Res. 2017b;26:225–233. Guo et al., 2018 Guo P, Zhang Y, Zhao Y. Biocapture of CO2 by different microalgal-based technologies for biogas upgrading and simultaneous biogas

slurry purification under various light intensities and photoperiods. Int J Environ Res Public Health. 2018;15(3):528. Ho et al., 2011 Ho S, Chen C, Lee D, Chang J. Perspectives on microalgal CO2emission mitigation systems e a review. Biotechnol Adv. 2011;29(2):189–198. Huang et al., 2017b Huang Q, Jiang F, Wang L, Yang C. Design of photobioreactors for mass cultivation of photosynthetic organisms. Engineering. 2017b;3:318–329. Jeong et al., 2013 Jeong H, Lee J, Cha M. Energy efficient growth control of microalgae using photobiological methods. Renew Energy. 2013;54:161–165. Jha et al., 2017 Jha D, Jain V, Sharma B, Kant A, Garlapati VK. Microalgaebased pharmaceuticals and nutraceuticals: an emerging field with immense market potential. Chem Bio Eng Rev. 2017;4(4):257–272. Juneja et al., 2013 Juneja A, Ceballos RM, Murthy GS. Effects of environmental factors and nutrient availability on the biochemical composition of algae for biofuels production: a review. Energies. 2013;6(9):4607–4638. Kao et al., 2012a Kao CY, Chiu SY, Huang TT, Dai L, Hsu LK, Lin CS. Ability of a mutant strain of the microalga Chlorella sp to capture carbon dioxide for biogas upgrading. Appl Energy. 2012a;93:176–183. Kao et al., 2012b Kao CY, Chiu SY, Huang TT, et al. A mutant strain of microalga Chlorella sp for the carbon dioxide capture from biogas. Biomass Bioenergy. 2012b;36:132–140. Kim et al., 2013 Kim TH, Lee Y, Han SH, Hwang SJ. The effects of wavelength and wavelength mixing ratios on microalgae growth and nitrogen, phosphorus removal using Scenedesmus sp for wastewater treatment. Bioresour Technol. 2013;130:75–80. Kougias and Angelidaki, 2018 Kougias PG, Angelidaki I. Biogas and its opportunities—a review. Front Environ Sci Eng. 2018;12(3):14. Kumari et al., 2009 Kumari A, Mahapatra P, Garlapati VK, Banerjee R. Enzymatic transesterification of Jatropha oil. Biotechnol Biofuels. 2009;2(1):1.

Kurano and Miyachi, 2005 Kurano N, Miyachi S. Selection of microalgal growth model for describing specific growth rate-light response using extended information criterion. J Biosci Bioeng. 2005;100(4):403–408. Lebrero et al., 2016 Lebrero R, Toledo-Cervantes A, Muñoz R, del Nery V, Foresti E. Biogas upgrading from vinasse digesters: a comparison between an anoxic biotrickling filter and an algal-bacterial photobioreactor. J Chem Technol Biotechnol. 2016;91(9):2488–2495. Lee et al., 2015 Lee E, Jalalizadeh M, Zhang Q. Growth kinetic models for microalgae cultivation: a review. Algal Res. 2015;12:497–512. López et al., 2018 López LR, Brito J, Mora M, et al. Feedforward control application in aerobic and anoxic biotrickling filters for H2S removal from biogas. J Chem Technol Biotechnol. 2018;93:2307–2315. Mandeno et al., 2005 Mandeno G, Craggs R, Tanner C, Sukias J, Webster-Brown J. Potential biogas scrubbing using a high rate pond. Water Sci Technol. 2005;51(12):253–256. Mandova et al., 2018 Mandova H, Leduc S, Wang C, et al. Possibilities for CO2 emission reduction using biomass in European integrated steel plants. Biomass Bioenergy. 2018;115:231–243. Mann et al., 2009 Mann G, Schlegel M, Schumann R, Sakalauskas A. Biogasconditioning with microalgae. Agron Res. 2009;7(1):33–38. Marín et al., 2019 Marín D, Ortíz A, Díez-Montero R, et al. Influence of liquidto-biogas ratio and alkalinity on the biogas upgrading performance in a demo scale algal-bacterial photobioreactor. Bioresour Technol. 2019;280:112–117. Markou et al., 2014 Markou G, Vandamme D, Muylaert K. Microalgal and cyanobacterial cultivation: the supply of nutrients. Water Res. 2014;65(C):186– 202. Meier et al., 2015 Meier L, Perez r, Azocar L, Rivas M, Jeison D. Photosynthetic CO2 uptake by microalgae: an attractive tool for biogas upgrading. Biomass Bioenergy. 2015;73:102–109. Meier et al., 2017 Meier L, Barros P, Torres A, Vilchez C, Jeison D.

Photosynthetic biogas upgrading using microalgae: effect of light/dark photoperiod. Renew Energy. 2017;106:17–23. Mera et al., 2016 Mera R, Torres E, Abalde J. Influence of sulphate on the reduction of cium toxicity in the microalga Chlamydomonas moewusii. Ecotoxicol Environ Saf. 2016;128:236–245. Mes et al., 2003 Mes TZD, de Stams AJM, Reith JH, Zeeman G. Methane production by anaerobic digestion of wastewater and solid wastes. In: Reith JH, Wijffels RH, Barten H, eds. Bio-Methane and Bio-Hydrogen: Status and Perspectives of Biological Methane and Hydrogen Production. The Hague, The Netherlands: Dutch Biological Hydrogen Foundation - NOVEM; 2003;58–102. Mohlin et al., 2018 Mohlin K, Camuzeaux JR, Muller A, Schneider M, Wagner G. Factoring in the forgotten role of renewables in CO2 emission trends using decomposition analysis. Energy Policy. 2018;116:290–296. Moutinho et al., 2018 Moutinho V, Madaleno M, Inglesi-Lotz R, Dogan E. Factors affecting CO2 emissions in top countries on renewable energies: a LMDI decomposition application. Renew Sustain Energy Rev. 2018;90:605–622. Muñoz and Guieysse, 2006 Muñoz R, Guieysse B. Algal-bacterial processes for the treatment of hazardous contaminants: a review. Water Res. 2006;40(15):2799–2815. Muñoz et al., 2015 Muñoz R, Meier L, Diaz I, Jeison D. A review on the stateof-the-art of physical/chemical and biological technologies for biogas upgrading. Rev Environ Sci Bio/Technol. 2015;14(4):727–759. Nag Dasgupta et al., 2010 Nag Dasgupta C, Gilbert JJ, Lindblad P, et al. Recent trends on the development of photobiological processes and photobioreactors for the improvement of hydrogen production. Int J Hydrog Energy. 2010;35(19):10218–10238. Nagarajan et al., 2019a Nagarajan D, Lee DJ, Chang JS. Biogas upgrading by microalgae: strategies and future perspectives. In: Alam M, Wang Z, eds. Microalgae Biotechnology for Development of Biofuel and Wastewater Treatment. Singapore: Springer; 2019a;347–395. Nagarajan et al., 2019b Nagarajan D, Lee DJ, Chang JS. Integration of anaerobic

digestion and microalgal cultivation for digestate bioremediation and biogas upgrading. Bioresour Technol. 2019b;290:121804. Nolla-Ardèvol et al., 2015 Nolla-Ardèvol V, Strous M, Tegetmeyer HE. Anaerobic digestion of the microalga Spirulina at extreme alkaline conditions: biogas production, metagenome, and metatranscriptome. Front Microbiol. 2015;6:597. Ouyang et al., 2015 Ouyang Y, Zhao Y, Sun S, Hu C, Ping L. Effect of light intensity on the capability of different microalgae species for simultaneous biogas upgrading and biogas slurry nutrient reduction. Int Biodeterior Biodegrad. 2015;104:157–163. Patterson et al., 2011 Patterson T, Esteves S, Dinsdale R, Guwy A. Life cycle assessment of biogas infrastructure options on a regional scale. Bioresour Technol. 2011;102(15):7313–7323. Pilon et al., 2011 Pilon L, Berberoğlu H, Kandilian R. Radiation transfer in photobiological carbon dioxide fixation and fuel production by microalgae. J Quant Spectrosc Radiat Transf. 2011;112(17):2639–2660. Posadas et al., 2017 Posadas E, Marín D, Blanco S, Lebrero R, Muñoz R. Simultaneous biogas upgrading and centrate treatment in an outdoors pilot scale high rate algal pond. Bioresour Technol. 2017;232:133–141. Prandini et al., 2016 Prandini JM, Da Silva MLB, Mezzari MP, Pirolli M, Michelon W, Soares HM. Enhancement of nutrient removal from swine wastewater digestate coupled to biogas purification by microalgae Scenedesmus spp. Bioresour Technol. 2016;202:67–75. Putt et al., 2011 Putt R, Singh M, Chinnasamy S, Das KC. An efficient system for carbonation of high-rate algae pond water to enhance CO2 mass transfer. Bioresour Technol. 2011;102(3):3240–3245. Rodero et al., 2018 Rodero MR, Ángeles R, Marín D, et al. Biogas purification and upgrading technologies. Biogas: Fundamentals, Process, and Operation Cham: Springer; 2018;239–276. Rodero et al., 2019a Rodero M, del R, Lebrero R, et al. Technology validation of photosynthetic biogas upgrading in a semi-industrial scale algal-bacterial

photobioreactor. Bioresour Technol. 2019a;279:43–49. Rodero et al., 2019b Rodero M, del R, Carvajal A, et al. Development of a control strategy to cope with biogas flowrate variations during photosynthetic biogas upgrading. Biomass Bioenergy. 2019b;131:105414. Ryckebosch et al., 2011 Ryckebosch E, Drouillon M, Vervaeren H. Techniques for transformation of biogas to biomethane. Biomass Bioenergy. 2011;35(5):1633–1645. Samudrala et al., 2017 Samudrala PJK, Garlapati VK, Dash A, Banerjee R, Scholz P. Sustainable green solvents and techniques for lipid extraction from microalgae: a review. Algal Res. 2017;21:138–147. Schmid et al., 2019 Schmid C, Horschig T, Pfeiffer A, Szarka N, Thrän D. Biogas upgrading: a review of national biomethane strategies and policies in selected countries. Energies. 2019;12(19):3803. Serejo et al., 2015 Serejo ML, Posadas E, Boncz MA, Blanco S, García-Encina P, Muñoz R. Influence of biogas flow rate on biomass composition during the optimization of biogas upgrading in microalgal-bacterial processes. Environ Sci Technol. 2015;49(5):3228–3236. Sevda et al., 2019 Sevda S, Garlapati VK, Sharma S, et al. Microalgae at niches of bioelectrochemical systems: a new platform for sustainable energy production coupled industrial effluent treatment. Bioresour Technol Rep. 2019;7:100290. Sevda et al., 2020 Sevda S, Garlapati VK, Datta P, et al. Bioethanol production from lignocellulosic/algal biomass:potential sustainable approach. In: Kothari R, Pathak VV, Tyagi VV, eds. Algal Biofuel: Sustainable Solution. New Delhi: The Energy and Resources Institute (TERI) Press; 2020;107–119. Sevda et al., 2021 Sevda S, Garlapati VK, Sharma S, Sreekrishnan TR. Potential of high energy compounds: hythane production. In: Singh L, Mahapatra DM, eds. Delivering Low-Carbon Biofuels With Bioproduct Recovery. New York: Elsevier; 2021;165–176. Sharma and Garlapati, 2021 Sharma S, Garlapati VK. Phycoremediation of Xray developer solution towards silver removal with concomitant lipid production. Environ Pollut. 2021;268A:115837.

Sharma et al., 2019 Sharma S, Gyeltshen T, Sevda S, Garlapati VK. Microalgae in bioelectrochemical systems: technological interventions. In: NavaniethaKrishnaraj R, Sani R, eds. Biovalorisation of Wastes to Renewable Chemicals and Biofuels. New York: Elsevier; 2019;361–371. Shen et al., 2020 Shen X, Xue Z, Sun L, et al. Effect of GR24 concentrations on biogas upgrade and nutrient removal by microalgae-based technology. Bioresour Technol. 2020;312:123563. Sierra et al., 2008 Sierra E, Acien FG, Fernandez JM, Garcia JL, Gonzalez C, Molina E. Characterization of a flat plate photobioreactor for the production of microalgae. Chem Eng J. 2008;138:136–147. Song et al., 2018 Song C, Fan Z, Li R, Liu Q, Kitamura Y. Efficient biogas upgrading by a novel membrane-cryogenic hybrid process: experiment and simulation study. J Membr Sci. 2018;565:194–202. Srinuanpan et al., 2017 Srinuanpan S, Cheirsilp B, Kitcha W, Prasertsan P. Strategies to improve methane content in biogas by cultivation of oleaginous microalgae and the evaluation of fuel properties of the microalgal lipids. Renew Energy. 2017;113:1229–1241. Srinuanpan et al., 2018 Srinuanpan S, Cheirsilp B, Prasertsan P. Effective biogas upgrading and production of biodiesel feedstocks by strategic cultivation of oleaginous microalgae. Energy. 2018;148:766–774. Srinuanpan et al., 2020 Srinuanpan S, Cheirsilp B, Kassim MA. Oleaginous microalgae cultivation for biogas upgrading and phytoremediation of wastewater. In: Yousuf A, ed. Microalgae Cultivation for Biofuels Production. Elsevier 2020;69–82. Su et al., 2017 Su Y, Song K, Zhang P, Su Y, Cheng J, Chen X. Progress of microalgae biofuel’s commercialization. Renew Sustain Energy Rev. 2017;74:402–411. Sun et al., 2016 Sun S, Ge Z, Zhao Y, Hu C, Zhang H, Ping L. Performance of CO2 concentrations on nutrient removal and biogas upgrading by integrating microalgal strains cultivation with activated sludge. Energy. 2016;97:229–237. Tan et al., 2016 Tan XB, Zhang YL, Yang LB, Chu HQ, Guo J. Outdoor cultures

of Chlorella pyrenoidosa in the effluent of anaerobically digested activated sludge: the effects of pH and free ammonia. Bioresour Technol. 2016;200:606– 615. Tang et al., 2011 Tang D, Han W, Li P, Miao X, Zhong J. CO2 biofixation and fatty acid composition of Scenedesmus obliquus and Chlorella pyrenoidosa in response to different CO2 levels. Bioresour Technol. 2011;102(3):3071–3076. Toledo-Cervantes et al., 2016 Toledo-Cervantes A, Serejo ML, Blanco S, Pérez R, Lebrero R, Muñoz R. Photosynthetic biogas upgrading to bio-methane: boosting nutrient recovery via biomass productivity control. Algal Res. 2016;17:46–52. Toledo-Cervantes et al., 2017a Toledo-Cervantes A, Lebrero R, Cavinato C, Muñoz R. Biogas upgrading using algal-bacterial processes. Microalgae-Based Biofuels and Bioproducts: From Feedstock Cultivation to End-Products Elsevier, Woodhead Publishing 2017a;283–304. Toledo-Cervantes et al., 2017b Toledo-Cervantes A, Madrid-Chirinos C, Cantera S, Lebrero R, Muñoz R. Influence of the gas-liquid flow configuration in the absorption column on photosynthetic biogas upgrading in algal-bacterial photobioreactors. Bioresour Technol. 2017b;225:336–342. Tongprawhan et al., 2014 Tongprawhan W, Srinuanpan S, Cheirsilp B. Biocapture of CO2 from biogas by oleaginous microalgae for improving methane content and simultaneously producing lipid. Bioresour Technol. 2014;170:90–99. Torres et al., 2020 Torres RÁ, Marín D, del Rosario Rodero M, et al. Biogas treatment for H2S, CO2, and other contaminants removal. In: Soreanu G, Dumont É, eds. From Biofiltration to Promising Options in Gaseous Fluxes Biotreatment. Elsevier 2020;153–176. Turon et al., 2016 Turon V, Trably E, Fouilland E, Steyer JP. Potentialities of dark fermentation effluents as substrates for microalgae growth: a review. Process Biochem. 2016;51(11):1843–1854. Vasumathi et al., 2012 Vasumathi KK, Premalatha M, Subramanian P. Parameters influencing the design of photobioreactor for the growth of microalgae. Renew Sustain Energy Rev. 2012;16(7):5443–5450.

Vo et al., 2019 Vo HNP, Ngo HH, Guo W, et al. A critical review on designs and applications of microalgae-based photobioreactors for pollutants treatment. Sci Total Environ. 2019;651:1549–1568. Wang et al., 2014 Wang XW, Liang JR, Luo CS, Chen , Gao YH. Biomass, total lipid production, and fatty acid composition of the marine diatom Chaetoceros muelleri in response to different CO2 levels. Bioresour Technol. 2014;161:124–130. Wang et al., 2017a Wang X, Bao K, Cao W, Zhao Y, Hu CW. Screening of microalgae for integral biogas slurry nutrient removal and biogas upgrading by different microalgae cultivation technology. Sci Rep. 2017a;7(1):1–12. Wang et al., 2017b Wang X, Gao S, Zhang Y, Zhao Y, Cao W. Performance of different microalgae-based technologies in biogas slurry nutrient removal and biogas upgrading in response to various initial CO2 concentration and mixed light-emitting diode light wavelength treatments. J Clean Prod. 2017b;166:408– 416. Wang et al., 2016 Wang Z, Zhao Y, Ge Z, Zhang H, Sun S. Selection of microalgae for simultaneous biogas upgrading and biogas slurry nutrient reduction under various photoperiods. J Chem Technol Biotechnol. 2016;91(7):1982–1989. Widjaja et al., 2009 Widjaja A, Chien CC, Ju YH. Study of increasing lipid production from fresh water microalgae Chlorella vulgaris. J Taiwan Inst Chem Eng. 2009;40(1):13–20. Xu et al., 2015 Xu J, Zhao Y, Zhao G, Zhang H. Nutrient removal and biogas upgrading by integrating freshwater algae cultivation with piggery anaerobic digestate liquid treatment. Appl Microbiol Biotechnol. 2015;99(15):6493–6501. Yan and Zheng, 2013 Yan C, Zheng Z. Performance of photoperiod and light intensity on biogas upgrade and biogas effluent nutrient reduction by the microalgae Chlorella sp. Bioresour Technol. 2013;139:292–299. Yan and Zheng, 2014 Yan C, Zheng Z. Performance of mixed LED light wavelengths on biogas upgrade and biogas fluid removal by microalga Chlorella sp. Appl Energy. 2014;113:1008–1014.

Yan et al., 2013 Yan C, Zhang L, Luo X, Zheng Z. Effects of various LED light wavelengths and intensities on the performance of purifying synthetic domestic sewage by microalgae at different influent C/N ratios. Ecol Eng. 2013;51:24–32. Yan et al., 2014 Yan C, Zhang L, Luo X, Zheng Z. Influence of influent methane concentration on biogas upgrading and biogas slurry purification under various LED (light-emitting diode) light wavelengths using Chlorella sp. Energy. 2014;69:419–426. Yan et al., 2016a Yan C, Zhu L, Wang Y. Photosynthetic CO2 uptake by microalgae for biogas upgrading and simultaneously biogas slurry decontamination by using of microalgae photobioreactor under various light wavelengths, light intensities, and photoperiods. Appl Energy. 2016a;178:9–18. Yan et al., 2016b Yan C, Muñoz R, Zhu L, Wang Y. The effects of various LED (light emitting diode) lighting strategies on simultaneous biogas upgrading and biogas slurry nutrient reduction by using of microalgae Chlorella sp. Energy. 2016b;106:554–561. Yeesang and Cheirsilp, 2011 Yeesang C, Cheirsilp B. Effect of nitrogen, salt, and iron content in the growth medium and light intensity on lipid production by microalgae isolated from freshwater sources in Thailand. Bioresour Technol. 2011;102(3):3034–3040. Zhang et al., 2017 Zhang Y, Bao K, Wang J, Zhao Y, Hu C. Performance of mixed LED light wavelengths on nutrient removal and biogas upgrading by different microalgal-based treatment technologies. Energy. 2017;130:392–401. Zhao et al., 2013 Zhao Y, Wang J, Zhang H, Yan C, Zhang Y. Effects of various LED light wavelengths and intensities on microalgae-based simultaneous biogas upgrading and digestate nutrient reduction process. Bioresour Technol. 2013;136:461–468. Zhao et al., 2015 Zhao Y, Sun S, Hu C, Zhang H, Xu J, Ping L. Performance of three microalgal strains in biogas slurry purification and biogas upgrade in response to various mixed light-emitting diode light wavelengths. Bioresour Technol. 2015;187:338–345. Zhao et al., 2019 Zhao Y, Guo G, Sun S, Hu C, Liu J. Co-pelletization of microalgae and fungi for efficient nutrient purification and biogas upgrading.

Bioresour Technol. 2019;289:121656.

Part IV Policy implications for biogas upgrading

Outline

Chapter 15 Biogas upgrading and life cycle assessment of different biogas upgrading technologies

Chapter 16 The role of techno-economic implications and governmental policies in accelerating the promotion of biomethane technologies

Chapter 17 Large-scale biogas upgrading plants: future prospective and technical challenges

Chapter 15

Biogas upgrading and life cycle assessment of different biogas upgrading technologies

Moonmoon Hiloidhari¹ and Shilpi Kumari², ¹1IDP in Climate Studies, Indian Institute of Technology Bombay, Mumbai, India, ²2Centre for Energy Studies, Indian Institute of Technology Delhi, New Delhi, India

Abstract

Biogas is an increasingly popular renewable energy source. It can be upgraded to biomethane for applications equivalent to fossil natural gas. The CO2 and other impurities in biogas, however, have to be removed prior to its use as biomethane. Pressure swing adsorption, high-pressure water scrubbing, membrane separation, and cryogenic separation are commercially available biogas upgrading technologies. The environmental performance of biogas upgrading can be assessed via life cycle assessment (LCA). Published LCAs of biogas upgrading technologies have mostly taken an attributional LCA approach with ReCiPe, CML-IA, and IMPACT 2002 as common choices for the impact assessment method. The studies are generally concerned with midpoint impacts associated with biogas upgrading (e.g., global warming potential, acidification, eutrophication, and ozone layer depletion). Including endpoint impact categories such as damage to human health and the ecosystem, and resource availability can further aid the decision maker. The consumption of electricity in the biogas upgrading phase, methane leakage, and heat demand in biogas production stage are the key contributors to the environmental impacts of biogas on biomethane upgrading technologies. Nevertheless, all the commercially available biogas upgrading technologies result in less greenhouse gas emissions than fossil-based natural gas production. Other impact categories like acidification and eutrophication show contrasting LCA results among the upgrading technologies. The environmental performance of biogas upgrading pathways can further be enhanced by (1) using renewable electricity at the upgrading phase, (2) reducing methane leakage from biogas production, storage, and transport, (3) using biogenic CO2 sources, (4) utilizing biogas digestates as organic nutrient, and (5) minimizing nitrogen fertilizer application for biomass raw material production.

Keywords

Organic waste; biogas; biogas upgrading; biomethane; LCA

Chapter outline

Outline


15.2 Biomethanation 415

15.2.1 Cleaning of biogas 416 15.2.2 Upgrading of biogas into biomethane 417

15.3 Brief overview of life cycle assessment 421

15.4 Life cycle assessment of biogas upgrading technologies 423


References 441

Further reading 445

15.1 Introduction

Biogas mainly consists of methane (CH4) (50%–75%), carbon dioxide (CO2) (25%–50%), hydrogen sulfides (H2S), hydrogen (H2), ammonia (NH3) (1%– 2%), and traces of other gases such as oxygen (O2) and nitrogen (N2) (Starr et al., 2012; Khan et al., 2017; Atelge et al., 2020). Biogas is a renewable and sustainable biofuel. It can reduce greenhouse gas emissions if well managed without leakage, and slurry available from biogas digester can be used as organic fertilizer (Atelge et al., 2020). Biogas can be produced from different types of organic biomass such as agricultural byproducts and residues, animal manure, and energy crops through an anaerobic digestion (AD) process (Montgomery and Bochmann, 2014; Bedoić et al., 2019; Wellinger and Murphy, 2013). In general, biogas production to the application includes feedstock selection and characterization, biogas generation using technologies like landfilling, AD, wastewater treatment, pretreatment of biogas to remove CO2 and other impurities to increase the methane content, injection of biomethane to the natural gas grid, or use as vehicular fuel (Starr et al., 2012). Various stages involved in biogas production are shown in Fig. 15.1 (Bedoić, et al., 2019; Masebinu et al., 2014; Bond and Templeton, 2011).

Figure 15.1 Various stages of biogas production and upgrading.

Biogas plants have been designed according to biological degradation processes that can be fully controlled and optimized to produce biogas. AD is a series of chemical reactions mediated by facultative bacteria and obligate bacteria which control CH4 production and also work at specific temperature zones: mesophilic (20°C–40°C) or thermophilic (above 40°C) (Bond and Templeton, 2011). AD is a controlled four-step process which involves (1) hydrolysis (conversion of complex carbohydrates, fats, and amino acids into simple monomers), (2) acidogenesis where glucose, amino acids, and long-chain fatty acids are converted into simple short chains, (3) acetogenesis, that is, production of hydrogen, carbon dioxide and acetic acids, and (4) methanogenesis (which involves production of biogas, that is, methane and carbon dioxide) (Rajendran et al., 2020). A general overview of biogas production from different types of biomass and AD is summarized in Fig. 15.2 (Meegoda et al., 2018; Dębowski et al., 2013; Rajendran et al., 2020).

Figure 15.2 Bacterial action in biomethane production from biomass (in Fig. 15.1). The flow chart represents complete pathways for biomethane production.

Cleaning, purifying, and upgrading of biogas is essential to utilize it as a modern industrial and vehicular fuel. There are a number of biogas upgrading technologies available, with each method having its own merits and demerits. Techniques currently matured and available at commercial markets include adsorption, absorption, membrane separation, and cryogenic. The performance, efficiency, energy, and materials input/output flows for these upgrading technologies vary, and consequently so do their environmental impact in of global warming potential (GWP) and other environmental indicators like eutrophication, acidification, ozone layer depletion, and energy consumption. The environmental performance of biogas production or upgrading technologies can be assessed through life cycle assessment (LCA). LCA is a tool that can be used to assess and compare the environmental impacts of different products or services throughout their entire life cycle, that is, from extraction of material, transport, production, use, and end-of-life treatment from raw materials extraction to production to consumption processes (European Commission, 2010; Lyng and Brekke, 2019). A number of LCA studies have been done on biogas production systems worldwide, generating useful information for policy makers and researchers to improve efficiency and decrease greenhouse gas (GHG) emissions and other environmental impacts (Hijazi et al., 2016). This chapter updates some of the recent developments in LCA of biogas upgrading technologies. The chapter is structured as follows: Section 15.1 provides a basic introduction to biogas, upgrading technologies, and LCA; Section 15.2 discusses the biomethanation process in detail; Section 15.3 gives an overview of the different stages involved in LCA; and Section 15.4 reviews some literature related to LCA of biogas upgrading technologies. The conclusions are discussed in the final section.

15.2 Biomethanation

In addition to CH4 and CO2 as major components, biogas also contains water, hydrogen sulfide (H2S), nitrogen (N2), oxygen (O2), ammonia (NH3), siloxanes, and particles (Masebinu et al., 2014). These impurities cause corrosion and low energy content of biogas; therefore it requires removal and cleaning of biogas. Biomethanation is a process of enrichment of CH4 by removing CO2 and other contaminants from biogas, making it suitable for injecting into the natural gas grid (Lorenzi et al., 2019). Different technologies are currently available, which remove CO2 from biogas to obtain a CH4-rich gas (>97%) (Adnan et al., 2019). Higher yields can be obtained through processes that recycle the CO2 into CH4. Removal of all impurities and CO2 can be divided into a two-step process: (1) cleaning of biogas and (2) upgradation of biogas into biomethane.

15.2.1 Cleaning of biogas

This involves the removal of impurities other than CO2 such as water, H2S, O2, N2, siloxane, and particulate matter.

15.2.1.1 Removal of water

Biogas in the digester is rich in water vapor, which condensates in pipelines and causes corrosion. Removal of water can be done by cooling, compression, absorption, or adsorption by increasing the pressure or decreasing the temperature (Petersson and Wellinger, 2009). Cooling is done by burying the gas

line equipped with a condensate trap in the soil. SiO2, activated carbon, or molecular sieves are used for adsorption (Petersson and Wellinger, 2009). These materials are usually regenerated by heating or a decrease in pressure. Absorption in glycol solutions or the use of hygroscopic salts are also used for water removal.

15.2.1.2 Removal of H2S

H2S can be removed by the addition of iron chloride to the digester slurry to precipitate out as iron sulfide and be removed together with the digestate (Persson et al., 2006; Ryckebosch et al., 2011). Adsorption on activated carbon can also remove H2S. Activated carbon with potassium iodide, potassium carbonate (K2CO3), or zinc oxide acts as a catalyst for H2S removal from biogas (Masebinu et al., 2014). H2S is also removed by biological scrubbing, that is, oxidation by chemoautotrophic microorganisms of the species Thiobacillus or Sulfolobus (Miltner et al., 2017).

15.2.1.3 Removal of other impurities

Other impurities present in biogas include O2, N2, NH3, siloxane, and particles. Particles in biogas can be removed by ing biogas over mechanical filters. O2, and N2 can be removed by adsorption with activated carbon, a carbon molecular sieve, or a membrane (Persson et al., 2006). These gases can also be removed by the removal process of CO2 from biogas in desulfurization processes. Some components of biogas which contain a silicon–oxygen bond are known as siloxanes. Removal of siloxane is mediated by cooling the gas, by adsorption on activated carbon, activated aluminum, or silica gel, or by absorption in liquid mixtures of hydrocarbons. It can also remove H2S during the cleaning process (Petersson and Wellinger 2009). Particulates which may cause mechanical wear in gas engines and gas turbines can be removed using

mechanical filters. Ammonia is also produced from the biomass, which is rich in protein and can be removed when the gas is dried or during the biogas upgrading process and NH3 removal does not require a separate process.

15.2.2 Upgrading of biogas into biomethane

The biomethanation process improves the quality of biogas fuel. The four main processes of biogas upgrading are adsorption, absorption, permeable membrane separation, and cryogenic separation. Adsorption and absorption technology are more efficient than the other two (Chen et al., 2015). The latter two technologies are high in cost and also in an early stage of development (Chen et al., 2015). Upgraded biogas, when used as fuel, has reduced GHG emissions and a less negative climatic impact.

15.2.2.1 Absorption

The separation of CO2 by absorption is based on different solubilities of various gas components in a liquid scrubbing solution. The solubility of gas CO2 is greater than that of CH4 in the applied liquid and thus the gas leaving the column has a greater concentration of CH4, while liquid in the column has a higher CO2 concertation (Persson et al., 2006). When water is used as a solvent for absorption, it is known as water scrubbing (Bauer et al., 2013). The technology of water scrubbing is used for biogas upgrading, which uses water as a solvent as the solubility of CO2 in water is higher than that of CH4. Similar to water scrubbing, organic solvent solution, that is, polyethylene glycol, is also used for the removal of CO2 (Nie et al., 2013). Chemical absorption such as amine scrubbing is used for CO2 removal (Bauer et al., 2013). This process is characterized by physical absorption of the biogas component in scrubbing liquid, followed by a chemical reaction between the liquid solvent and gas component. The chemical reaction is highly selective and the solubility of CH4

is very low, resulting in high CH4 recovery. CO2 has high affinity for solvents [mainly aqueous solutions of monoethanolamine (MEA), diethanolamine (DEA), and methyldiethanolamine (MDEA)], causing separation of CO2 and enrichment of CH4 (Nie et al., 2013).

15.2.2.2 Adsorption

Adsorption is based on the adsorption of various gas components of biogas on a solid surface under high pressure. Pressure swing absorption (PSA) is one of the best known absorption processes for the removal of CO2 (Grande, 2012). Adsorption of CO2 on a solid surface and its removal by changing pressure (high to low and low to high) is known as pressure swing adsorption. CO2 is separated by adsorption at high pressure and the resultant gas is rich in CH4 (Masebinu et al., 2014). This is followed by decreasing the pressure and adsorbing material is regenerated for the next sequence of reloading. PSA may have four, six, or nine adsorber vessels in parallel at different positions so that the process operates continuously (Masebinu et al., 2014). The gas that is desorbed during the first and eventually the second pressure drop may be returned to the inlet of the raw gas since it will contain some CH4 that was adsorbed together with carbon dioxide. The gas desorbed in the following pressure reduction step is either led to the next column, or if it is almost entirely CH4 free, it is released to the atmosphere. Also, this parallel arrangement of vessels allows entry of gas to the next vessels if it contains a small amount of CH4, which may be adsorbed together with CO2 in the first vessel or be released to the atmosphere if it is completely free from CH4.

15.2.2.3 Membrane separation

This is based on the principle that it is permeable for CO2, water, NH3, O2, and N2, and not permeable to CH4. Typical membranes for biogas upgrading are

made of polymeric materials like polysulfone, polyimide, or polydimethylsiloxane (Molino et al., 2013; Sridhar et al., 2007). Membrane technology was used to separate CO2 from biogas in order to obtain biomethane of suitable quality for placing it into the national distribution network. The first membrane module bifurcates the feed gas into biomethane (>95%.vol) at high pressure (30 bar) and permeates gas rich in impurities including CO2, N2, O2, H2, water, and additional impurities collected at a low pressure of 2 bar (Molino et al., 2013). The permeate gas is again circulated to recover CH4 (if present) by more than 85%.vol. The permeate gas generated after the second membrane is vented as it contains impurities (Molino et al., 2013). The membrane also removes acid gases and water vapor to improve gas quality.

15.2.2.4 Cryogenic separation

Cryogenic separation uses the difference of temperature for separation of different gas species (Awe et al., 2017). CO2 has a desublimation temperature of −78.5°C at atmospheric pressure, while CH4 condenses at −161°C (Jonsson and Westman, 2011). CO2 is separated from CH4 at low temperature with continuous sublimation (gas to solid) and desublimation of CO2. The first step is compression of raw biogas to 17–26 bar and then cooled to −26°C for removal of H2S, SO2, halogens, and siloxane (Jonsson and Westman, 2011; Petersson and Wellinger 2009). When CO2 in biogas is desublimated it follows that the partial pressure of CO2 is reduced, therefore the concentration of CO2 is lowered and a lower temperature will be required to further desublimate the CO2. A lower temperature results in a higher removal efficiency of carbon dioxide (Jonsson and Westman, 2011). However, the presence of CH4 in the biogas mixture affects the characteristics of the gas, thus requiring higher pressure and/or lower temperature to condense CO2. To avoid freezing and other problems in the cryogenic process, water and H2S need to be removed before cryogenic separation of CO2 and CH4 (Masebinu et al., 2014). A summary of the advantages and disadvantages of the commercially available biogas upgrading technologies is presented in Table 15.1.

Table 15.1

Techniques

Advantages

Absorption

• Highly efficient (97%–99%) CH4 with very less loss of (0/1%–2%) of CH4 • Simultane

Adsorption

• Highly efficient (98% CH4) • May be applicable for small to medium plants • Adsorben

Membrane separation

• In comparison to above, it requires less investment • Operation is simple and high reliab

Cryogenic technologies

• High CH4 (90%–98%) • Low extra energy cost to reach liquid biomethane

The review suggests that water scrubbing technology is preferable to the PSA method and chemical absorption. Water scrubbing, and absorption with chemical or physical solvents are conventional technologies in biogas upgradation. The technology should be process-technology feasible and energy-economyenvironmentally favorable. From a process-technology perspective, specifications like CH4 purity level and flow rate range were evaluated. Energy and economy-environmental favorable technology should be evaluated on the basis of the assessment of consistency of available data and on utilization of the data for comparative analysis of given technology (Nie et al., 2013). The above section discussed the upgrading of biogas to biomethane through sorption (physical and chemical) or separation (cryogenic or membrane) techniques. One of the major issues in biogas upgradation is the high cost involved in obtaining methane-rich biogas. Several new technologies are under different stages of development which can reduce the cost of upgrading, increase the CH4 purity, and also reduce the waste discharge. Some of them are:

1. In situ CH4 enrichment: In situ methane enrichment is a solution to reduce the cost of existing commercial biogas upgradation technologies. This is based on the recirculation of liquid sludge from the digestion chamber to the desorption column where it undergoes counterflow of O2 and N2 by the which CO2 is dissolved in the sludge. This again is circulated for additional absorption of CO2 (Kadam and Panwar, 2017). 2. Hybrid technologies: These technologies combine two different techniques, such as pressurized water scrubbing, amine absorption, and cryogenic separation, to develop novel hybrid membrane processes. Combining two or more technologies not only improves the upgradation process but also increases the profit for commercial use (Scholz et al., 2013). 3. Industrial lung: This is a biotechnologically hybridized process using carbonic anhydrase enzyme which enhances and catalyzes the breakdown of CO2 formed from cell metabolism (Scholwin and Held, 2013).

4. Supersonic separation: Supersonic separation is a recent method comprising of a compact tubular device that efficiently combines expansion, cyclonic gas/liquid separation, and recompression using a convergent–divergent nozzle. The nozzle is used to expand the feed biogas to supersonic velocity, which causes a drop in temperature and pressure (Sahota et al., 2018). This leads to mist droplets of condensed water and hydrocarbon, and simultaneously condenses and separates water and hydrocarbons from biogas. 5. CO2 utilization: This is used for enrichment of CH4 by CO2 utilization by (A) a chemical process or (B) a biological process (Adnan et al., 2019). In a biological process, microorganisms convert CO2 into useful products, that is, biological fixation (Wu et al., 2017). Meanwhile in a chemical process, biogas is used directly as feedstock for CO2 methanation into either CH4 or CO (Adnan et al., 2019). 6. CO2 conversion into succinic acid (SA) (C4H6O4): This involves bacterial fermentation using Actinobacillus succinogenes while also simultaneously producing high-purity CH4 (Gunnarsson et al., 2014), and the produced SA can be sold to the market which reduces the operational cost. Therefore, this technique could be further explored and improved. Research work is on-going to improve the biogas upgradation and enrichment of CH4 and also to reduce waste release into the atmosphere.

15.3 Brief overview of life cycle assessment

LCA is a computational process to assess the environmental performance of a product, system, or service considering the entire life cycle. The application of LCA is wide-ranging, for example, environmental assessment and planning (Arvidsson et al., 2018; Zhang et al., 2020), to agriculture (Liang et al., 2019), forestry (Costa et al., 2018), renewable and nonrenewable energy (Petrillo et al., 2016; Lelek and Kulczycka, 2020), product rating, and eco-friendly design (Marconi and Favi, 2020). It helps to understand and improve the environmental performance of a product or service, including energy, and carbon and water footprints. An LCA study should follow an ISO standard such as 14040/44 comprising the four major stages: goal and scope definition, life cycle inventory (LCI), life cycle impact assessment (LCIA), and interpretation (Guinée, 2001; ISO, 2006). An LCA could be attributional, consequential, or a combination of both (hybrid LCA). In an attributional LCA (ALCA), the impact of the processes required to produce, consume, or dispose a product is assessed. However, it does not consider what is happening outside the system, that is, impact/changes outside the studied system boundary (Ekvall, 2020). In contrast, consequential LCA (CLCA) attempts to understand the consequences or level of modifications caused by a product outside its system boundary (Ekvall, 2020). In some cases, both attributional and consequential LCAs are applied together in a hybrid mode. A common example of the application of a hybrid LCA is observed in the case of studies related to biomass-based renewable electricity and ethanol or biodiesel (Wang et al., 2020). In some biomass energy systems (e.g., sugarcane-based electricity and ethanol) apart from the main product (sugar), a number of coproducts (bagasse, molasses, press mud) and useful wastes (sugarcane leaves and tops, bagasse ash) are generated, requiring not only allocation of environmental burden between the product and coproducts but also system expansion to describe the potential uses of the waste. Thus allocation (part of ALCA) and system expansion (part of CLCA) can be equally used for such studies following a hybrid LCA approach. The first step of an LCA study is to define its goal and scope. Here, the purpose of the study, intended audience, system boundary, functional unit, allocation rule,

and assumptions are explained. This phase governs the subsequent stages of the LCA and also allows to consider which processes and flows should be included in or excluded from the study. The system boundary defines the start- and endpoints of an LCA study as well as the processes considered within the boundary. If some processes are not included within the system boundary, for example, machinery and infrastructure building for a power plant, it has to be clearly explained. In many LCAs (mostly common with biomass energy studies), machinery and infrastructure construction are excluded from the system boundary. Seabra et al. (2011) noted that the inclusion of the embodied energy of machinery and buildings in LCA is unnecessary as they represent small shares of total emissions and energy use. The functional unit explains the function of the product under investigation. It also enables comparison of two different systems. For example, in the case of LCA of biogas upgrading technologies, the functional unit could be the production or use of 1 m³ upgraded biogas (biomethane). In the case of biomass electricity, the functional unit could be the production of 1 kWh or 1 MJ of biomass electricity. When a system produces multiple outputs, for example, in sugar industries, sugar, molasses, press mud, and bagasse are produced, which can further be used for ethanol or electricity production or other value addition. Therefore in such cases, the resulting environmental burden has to be equally divided among the outputs (product and coproducts) through an allocation principle. The allocation is the partitioning of the environmental burden between the product and the coproduct(s) of a multifunctional process (Reap et al., 2008). Allocation can be physical (mass, energy, exergy) or economic. Sometimes system expansion is performed to avoid allocation. The ISO standard recommendation is to avoid allocation wherever possible (Cherubini, 2010). However, system expansion is complex, require knowledge of the situation inside and outside the system boundary. It is also influenced by market behavior, which is not easy to predict precisely always. So, when system expansion is not possible, LCA practitioners can opt for allocation-based ACLA. Nevertheless, the final outcomes of ALCA are greatly influenced by the allocation rule and therefore the choice of a particular allocation should be properly justified in the study. The second stage of an LCA is the LCI. At the LCI phase, all the relevant flows and processes within the studied systems are identified, recorded, and analyzed (Guinée, 2001; ISO, 2006). It takes into the upstream/downstream and

foreground/background processes. A comprehensive list of inventory is necessary for a detailed assessment of the environmental performance of the system. Available databases like EcoInvent, Agrofootprint, or published literature can be used for the inventory lists. However, a common drawback of too much reliance on the LCA database is that the inventories may not be available for some processes or even if available they may not accurately depict the situation of a particular region. For example, inventories for emissions associated with coal-based electricity production in a state of India are not the same as in Europe due to differences in the quality and heating value of coal. Therefore primary data-based (survey, experiment, and modeling) LCI building should be the priority as far as possible. The LCIA stage uses the information generated in the LCI phase to produce environmental impact results of a studied system. The impacts are classified, characterized (mandatory as per ISO 14040/44 rules), normalized, ranked, grouped, and weighted (optional stages) (ISO, 2006; Pizzol et al., 2017). The impact can be identified at midpoint or endpoint level (ISO, 2006). Each midpoint indicator focuses on a single environmental issue. Midpoint indicators are higher in number compared to the endpoint indicators. The midpoint indicators are linked to different damage pathways, which leads to endpoint indicators. A midpoint indicator assesses the impact of the cause–effect chain before reaching the endpoint. For example, GWP, ozone depletion, and human toxicity are midpoint indicators. Together they have an impact on human health, which is an endpoint indicator. There are standard methods available to conduct the LCIA, which can focus on a single issue (e.g., Cumulative Energy Demand, IPCC 2013 GWP 100a) or multiple issues (e.g., ReCiPe 2016, IMPACT 2002+, CML-IA). Again the LCIA can be region-specific. For example, TRACI is a midpoint-level LCIA tool developed by the USEPA, especially for North America (Bare, 2002). The latest ReCiPe 2016 impact assessment method covers a global perspective with 18 midpoint and three endpoint indicators. The final stage of the LCA is the interpretation. This stage makes conclusions and recommendations based on the LCI and LCIA. Uncertainty and sensitivity analysis are also done in this phase to understand the level of uncertainty associated with the LCA study and identify the sensitive parameters. Monte Carlo Simulation is generally done to conduct uncertainty and sensitivity analysis (Groen and Heijungs, 2017). A complete LCA study should include the uncertainty and sensitivity analysis (Agostini et al., 2020). LCAs without uncertainty and sensitivity analysis are not reliable and the outcomes may lead to

erroneous decision-making. Some software available for conducting LCA are listed in Table 15.2.

Table 15.2

Software

Developer

Source

SimaPro

PRé Sustainability

www.simapro.com/

GaBi

sphera

www.gabi-software.com/international/index/

Umberto

Ifu Humbarg

www.ifu.com/en/umberto/lca-software/

OpenLCA

GreenDelta

www.openlca.org/

EIO-LCA

Carnegie Mellon University

www.eiolca.net/

CMLCA*

Leiden University

www.cmlca.eu/

eBalance

IKE Environmental Technology Co. Ltd

www.ike-global.com/products-2/lca-software-ebalanc

EIME

Bureau Veritas CODDE

www.codde.fr/en/our-software/eime-presentation

EarthSmart

EarthShift Global

www.earthshiftglobal.com/software/earthsmart-lca-so

*This is in Dutch and CML translates to English as Institute of Environmental Sciences (CML, Centrum voor Milieukunde Leiden).

15.4 Life cycle assessment of biogas upgrading technologies

Biogas production and upgrading are two essential parts of the biomethane production chain. Knowing the environmental profile of biogas upgrading technologies is important and LCA-based studies enable identifying emissions hotspots and improving the efficiency upgrading technologies. Different stages of LCA of biogas upgrading are explained in Fig. 15.3.

Figure 15.3 Different stages of life cycle assessment of biogas upgrading. Reused with permission from Starr, K., Gabarrell, X., Villalba, G., Talens, L. and Lombardi, L., 2012. Life cycle assessment of biogas upgrading technologies. Waste Manage., 32 (5), pp. 991–999.

Some studies have focused only on LCA of biogas production (Wang et al., 2016; Aziz et al., 2019), while some have limited the scope to LCA of biogas upgrading technologies. In the next section, a review of published literature on LCA of biogas upgrading technologies is discussed. Peer-reviewed research articles are searched on Google Scholar, ScienceDirect, and SpringerLink using keywords “LCA of biogas upgrading technologies,” “LCA of biogas to biomethane,” “LCA of biomethane,” “LCA of biogas upgrading via adsorption,” “LCA of biogas upgrading via absorption,” “LCA of biogas upgrading via membrane separation,” and “LCA of biogas upgrading via cryogenic separation.” Both the title and abstract were first scanned to examine if the article contained any of the keyword(s), particularly LCA of at least one biogas upgrading technology. Articles that matched with the search criteria were selected for the next stage of scanning, that is, examining if the study followed standard LCA protocol. The selected articles are discussed below. Starr et al. (2012) conducted a comparative LCA of three different biogas upgrading technologies, namely high-pressure water scrubbing (HPWS), alkaline with regeneration (AwR), and bottom ash upgrading (BABIU). The LCA was done using GaBi software. The system boundary included biogas coming from landfill, then biomethane production (using the three different upgrading technologies and then injecting biomethane to the natural gas grid). The functional unit was the removal of 1 ton of CO2. The LCIA methods include CML 2001 and cumulative energy demand (CED). The CED assesses the total primary energy demand (both renewable and nonrenewable) to produce a product or service. The CML impact categories included are: (1) abiotic depletion, elements; (2) abiotic depletion, fossil; (3) acidification potential; (4) eutrophication potential; (5) freshwater aquatic ecotoxicity potential; (6) human toxicity potential; (7) marine aquatic ecotoxicity potential; (8) ozone layer depletion potential; (9) terrestrial ecotoxicity potential; (10) GWP; and (11) photochemical ozone creation potential. The authors find that the AwR upgrading technology had the highest environmental impacts due to the high

energy demand for alkaline reactants production. The BABIU technology was found to be most environmentally friendly when compared with different technologies such as HPWS, AwR, BABIU, organic physical scrubbing (OPS), PSA, MS (membrane separation), and CS (cryogenic separation). The electricity input varies from 314 to 1264 MJ per ton of CO2 removal among these technologies, with the highest being the MS and the lowest the AwR. Biomethane yield is the highest in the MS technologies, but the methane purity is lowest and the methane loss is also highest. The BABIU process had the lowest impact in most categories, even when compared to five other CO2 capture technologies on the market. AwR and BABIU have a particularly low impact in the GWP category as a result of the immediate storage of the CO2. For AwR, the authors determined that using NaOH instead of KOH improves its environmental performance by 34%. For the BABIU process the use of renewable energies would improve its impact since it s for 55% of the impact. Florio et al. (2019) conducted an LCA of different biogas upgrading technologies, namely MS, CS, PSA, chemical scrubbing, and HPWS, and compared them with a reference scenario of a combined heat and power (CHP) plant. The biogas was obtained from mixed biogenic waste feedstocks (solid and liquid manure, biowaste, sewage sludge, and used vegetable cooking oil). The LCA was attributional. The cradle-to-gate system boundary was limited to biogas production and conversion to biomethane, or into electricity or heat. The functional unit was the use of 1 m³ of biogas produced from biogenic waste feedstock. As an allocation rule, the authors applied a system expansion to credit the natural gas substitution by biomethane. The impact assessment was conducted using the ReCiPe 2016 midpoint (H) and the impact categories considered were (1) global warming, (2) stratospheric ozone depletion, (3) terrestrial acidification, (4) freshwater eutrophication, (5) human toxicity (both carcinogenic and noncarcinogenic), (6) mineral resource scarcity, (7) fossil resource scarcity, and (8) water consumption. The authors found that the membrane upgrading technology was slightly more environmentally beneficial than the other technologies. They also observed that the environmental gains could be further enhanced if the digestate (coproduct of the AD process) was used as organic fertilizer, thus avoiding the impacts of chemical fertilizers. Biogas digestate is rich in nitrogen, phosphorus, and potassium (NPK), and can replace inorganic fertilizer (Tambone et al., 2010; Walsh et al., 2012). Uncertainty analysis using Monte Carlo simulation shows that the impact categories of freshwater eutrophication, human toxicity, and water consumption

have higher uncertainty values, requiring further studies. Future improvement in of environmental advantages of CO2 recovery from upgrading technologies has been suggested by the authors (Florio et al., 2019). The importance of site-specific data coupled with industrial case studies are required to bridge the knowledge gap between laboratory tests and large-scale operations. The need for economic evaluation of biomethane is also suggested by the authors to aid the renewable energy decision-making process. Castellani et al. (2018) conducted an attributional LCA of hydrate-based biogas upgrading technology. The goal of the study was to assess primary energy consumption and GHG emissions associated with the production of biosynthetic methane via hydrate-based biogas upgrading and CO2 methanation. The system boundary included all the phases from the production of input materials, energy, chemicals, to the assembly of the final product, biosynthetic methane (a mixture of methane-biogas and a synthetic fuel). The IPCC 2013 GWP 100a and CED were taken as LCIA methods since the study was focused on only carbon and energy footprints. The authors reported that the carbon and energy footprints of biosynthetic methane are 0.7081 kg CO2eq and 28.55 MJ/Nm³. The hydrogen production phase contributes to the maximum environmental impact. Lorenzi et al. (2019) performed a comparative attributional LCA between solid oxide electrolyzer cells and HPWS biogas upgrading technology. The functional unit was the production of 1 m³ biomethane. The ReCiPe 2016 midpoint (H) method was used for impact assessment. The source of electricity for the upgrading process and carbon content of electricity are the key factors that determine the environmental burden of biogas upgrading. Increasing the share of low-carbon renewable electricity will significantly lower the environmental impact. The authors found that in an SOEC-based plant, when the share of renewable electricity drops below 80%, the environmental impact goes up in all impact categories, even though the methane yield is higher. Therefore a low- or zero-carbon electricity source for the upgrading process is essential for SOEC technology. Biogas production from algae biomass is also reported in the literature (Wiley et al., 2011; Dębowski et al., 2013; Uggetti et al., 2016). Ferreira et al. (2019) assessed life cycle energy consumption and CO2 emissions of pilot- and realscale photosynthetic algal biogas upgrading units. They conducted an attributional LCA taking MJ of CH4 produced as the functional unit. The LCA was gate to gate, being limited to the microalgae production system, biogas

upgrading, harvesting, and drying. Using the ReCiPe 2016 midpoint impact assessment, the study found that at an industrial scale, biomass dewatering and drying ed for the highest energy consumption, CO2 emissions, and environmental impacts, followed by the energy consumption for agitation of the algal-bacterial broth at HRAP (high-rate algal pond). It was also observed that the biogas upgrading unit shows high impacts on climate change and low effects on fossil and water depletion. Production of diesel fuel and gasoline via Fischer–Tropsch (FT) synthesis and a dry reforming process using biogas have been also reported in the literature (Okeke and Mani, 2017; Navas-Anguita et al., 2019). LCA studies can reveal the environmental performance of such a conversion path. In this regard, NavasAnguita et al. (2019) conducted an attributional LCA of synthetic fuel produced from FT synthesis and a dry reforming process. The suitability of produced synthetic biodiesel was compared with fossil diesel. The overall process of the production of synthetic fuel and electricity (as coproduct) from biogas can be grouped into (1) syngas production via dry reforming of methane, syngas cleaning and conditioning, hydrocarbon production through FT synthesis, refining of FT products, and power generation under a combined-cycle scheme (Navas-Anguita et al., 2019). The functional unit of the study was the production of 1 kg of synthetic biodiesel at the plant. The system boundary of the study is presented in Fig. 15.4. The impacts were assessed using the CML method. The impact categories considered are (1) global warming, (2) cumulative nonrenewable energy demand, (3) ozone layer depletion, (4) acidification, and (5) eutrophication. Since electricity is also produced along with fuels, the author used an economic allocation among diesel, gasoline, and electricity for environmental burden sharing. The authors recommended not using energybased allocation because it would greatly influence the environmental characterization results, misrepresenting the main function of the system (i.e., production of biofuels). When compared with fossil diesel, synthetic biodiesel underperforms in of global warming, acidification, and eutrophication. Therefore it was noted that a reduction in methane leakage, heat demand at the biogas production phase, and reducing NH3 and NO emissions in the biogas-toliquid plant are necessary to improve the overall environmental performance of the biogas to synthetic biodiesel system.

Figure 15.4 System boundary of life cycle assessment of synthetic biodiesel production from biogas. Reused with permission from Navas-Anguita, Z., Cruz, P.L., Martin-Gamboa, M., Iribarren, D., Dufour, J., 2019. Simulation and life cycle assessment of synthetic fuels produced via biogas dry reforming and Fischer-Tropsch synthesis. Fuel, 235, pp. 1492–1500.

Lombardi and Francini (2020) compared five biogas to biomethane upgrading technologies (HPWS, amine scrubbing, potassium carbonate scrubbing, PSA, and membrane permeation). The functional unit of the attributional LCA was the upgrading of 1 Nm³ of raw biogas. A municipal organic solid waste-based biogas production phase was also included in the system boundary. The impact assessment was done using the CML-IA baseline method. The authors observed that amine scrubbing proved to be the best performing upgrading technology in of GWP, mainly due to low methane leakage and minimal electricity consumption. The impact on the GWP indicator mostly comes from methane emissions in the case of PSA, high-pressure water scrubbing, and membrane permeation. The contribution of solvents to the GWP is significant in the case of amine scrubbing, but less in potassium carbonate scrubbing. However, PSA is better technology when it comes to a human toxicity indicator. In of economic analysis through life cycle costing, the authors noted that highpressure water scrubbing is most economical for small plants and PSA is economically most suitable when the plant is large. Ardolino et al. (2018) conducted an attributional LCA to assess the environmental performance of biomethane (upgraded via the MS process) as transportation fuel. The functional unit was the treatment of 100 t/d of organic waste and impact assessment method was IMPACT 2000+. The system boundary included all the activities from biowaste delivery at the plant to the management of all outputs (e.g., biomethane as a transportation fuel) and residues. Instead of allocation, the authors used system expansion to explain the avoided impact of fossil diesel through the utilization of biomethane as a vehicular fuel. The authors observed that the production of biomethane for road transportation is better than the production of energy in of a reduction in GHG emissions and nonrenewable energy consumption. Using GaBi software, Leonzio (2016) compared the environmental performance

of biogas upgrading by a chemical adsorption process using three different aqueous solutions of MEA (monoethanolamine), NaOH (sodium hydroxide), and KOH (potassium hydroxide). The LCA was attributional and the functional unit was the generation of 1 kWh of biomethane. For the materials and energy balances for the LCI, the authors used the ChemCad 6.3 simulations. Since this simulation does not produce a coproduct or byproduct, no allocation was applied. The study also did not consider methane leakage or fugitive emissions from the AD plant, pipes, valves, over-pressure of the system, and the storage facilities for waste and biogas, considering it is difficult to estimate these emissions due to their variability from one site to another (Møller et al., 2009). The impact assessment method was CML 2001. The LCA results reveal that the environmental impact of the process using KOH aqueous solution is lower, followed by chemical absorption with MEA and NaOH aqueous solution. Power-to-gas (P2G or PtG) is a relatively new concept where an excess amount of electricity, mostly from renewable sources like solar and wind is used to convert water into renewable hydrogen and oxygen via an electrolysis process. The oxygen is released into the atmosphere and the hydrogen is captured, mixed with natural gas, and stored in the pipeline system. In the biomethanation process, the renewable hydrogen is combined with CO2 to produce renewable methane gas (SoCalGas, 2020). Renewable hydrogen can also be used to produce synthesis gas, electricity, liquid fuels, or chemicals (Wulf et al., 2018). The P2G process is illustrated in Fig. 15.5.

Figure 15.5 Power-to-gas process. SoCalGas. Power-to-Gas Technology. 2020. https://www.socalgas.com/smart-energy/renewable-gas/power-to-gas (accessed 20.07.20).

Collet et al. (2017) conducted a technoeconomic and LCA of methane production using biogas upgrading and P2G technology. The functional unit was the production of 1 MJ energy through methane combustion in a boiler based on the lower heating value of methane (50 MJ/kg). Continuous P2G operation results in more GHG emissions than direct injection, and therefore intermittent operation using renewable electricity is suggested to reduce the emissions burden (Collet et al., 2017). It was also observed that continuous P2G leads to a higher environmental impact than biogas upgrading, but lower than fossil energy. As an improvement, lowering electricity consumption of the electrolysis process and integrating renewable credits from CO2 valorization are suggested to increase the competitiveness of the technology (Collet et al., 2017). Zhang et al. (2017) also conducted an LCA of the P2G process. Similar to Collet et al. (2017), the study also noted that the type of electricity (renewable, nonrenewable) and sources of CO2 greatly influence the life cycle GHG emissions of P2G. The electricity supply from fossil fuels results in P2G causing higher life cycle GHG emissions than conventional technologies (Zhang et al., 2017). Therefore to environmentally compete with conventional hydrogen and natural gas, the electricity supply from a low GHG-intensity source and CO2 supply from biogenic sources, including direct air capture, need to be ensured for P2G technology. When system expansion is applied, the authors found that power-to-hydrogen (P2H) replacing hydrogen from fossil sources can more effectively contribute to GHG emissions reductions than synthetic natural gas (SNG) replacing natural gas as a vehicle fuel. Nevertheless, the authors also concluded that considering the well-established existing natural gas grid and infrastructure, and the numerous options to use natural gas, large-scale conversion of electricity to SNG might be more realistic than large-scale P2H. From this limited review of selected literature on LCA of biogas upgrading technologies, it is observed that LCA is a powerful tool to understand the environmental performance of biogas upgrading technologies and to better manage the upgrading processes. The above LCA studies have noted that the

consumption of electricity in the biogas upgrading phase and leakage of methane in the biogas production phase are the two main contributors to the overall environmental impacts. Reagents and solvents using chemical-oriented biogas upgrading technologies are also responsible for the environmental impact categories. It is also observed that all the biogas to biomethane/bionatural gas upgrading technologies result in lower greenhouse gas emissions compared to fossil natural gas production. However, other impact categories, for example, acidification and eutrophication, do not show such common similarities among the upgrading technologies. P2G shows a promising application for the use of excess electricity derived from solar or wind to produce biomethane, diesel, or other chemical production. However, consumption of a high amount of fossil fuel-based electricity for P2G leads to higher GWP, therefore ensuring the supply of renewable electricity and biogenic CO2 needed for future improvement in P2G technology. A summary of the results of the reviewed LCA studies of biogas upgrading technologies is presented in Table 15.3.

Table 15.3

Author(s)

Biogas upgrading technology

Starr et al. (2012)

High-pressure water scrubbing (HPWS), alkaline with regeneration (AwR) and bot

Florio et al. (2019)

MS, CS, PSA, chemical scrubbing (CSc), HPWS

Castellani et al. (2018)

Hydrate-based biogas upgrading

Lorenzi et al. (2019)

Solid oxide electrolyzer cells (SOECs) and HPWS

Ferreira et al. (2019)

Photosynthetic algae biogas upgrading units

Navas-Anguita et al. (2019)

Fischer–Tropsch (FT) synthesis and dry reforming process

Lombardi and Francini (2020)

HPWS, amine scrubbing (AS), potassium carbonate scrubbing (PCS), PSA, and m

Ardolino et al. (2018)

MS

Leonzio (2016)

Chemical adsorption

Collet et al. (2017)

Power-to-gas (P2G)

Zhang et al. (2017)

P2G

15.5 Conclusions

Biogas is a renewable and sustainable fuel which not only results in less greenhouse gas emissions than fossil fuels but also provides organic fertilizer as a coproduct. Biogas can be upgraded to biomethane, a methane-rich energy source for applications equivalent to fossil natural gas. However, CO2 and other impurities have to be removed from raw biogas for its use as a vehicular fuel or natural gas. Biogas upgrading technologies that are currently matured and commercially available include adsorption, absorption, MS, and cryogenics. The system performance, efficiency, energy and materials flows, and environmental performance vary among the upgrading technologies. LCA studies of biogas upgrading technologies have reported that the consumption of electricity in the biogas upgrading phase and the leakage of methane in the biogas production phase are the two main contributors to the overall environmental impacts. The share of reagents and solvents contributes to the environmental impacts of chemical-based biogas upgrading technologies. Lower greenhouse gas emissions are reported for all the biogas upgrading technologies compared to fossil natural gas production except for P2G technology. Nevertheless, some impact categories such as acidification and eutrophication have varied performance among the upgrading technologies. The P2G technology can use excess electricity derived from solar or wind energy systems to produce biomethane, diesel, or other chemicals. Ensuring a renewable electricity supply and biogenic CO2 source are required as future improvements in P2G technology to lower the GHG emissions. Key future improvements required to further enhance the environmental performance of biomethane production and upgrading pathways include (1) increasing the share of zero- or low-carbon renewable electricity at the upgrading phase, (2) reducing methane leakage from biogas production, storage, and pipeline transfer stages, (3) proper management of digestates (e.g., as organic fertilizer), and (4) minimizing nitrogen fertilizer use in the biomass feedstock production/cultivation stage.

Acknowledgment

The first author is grateful to the Indian Institute of Technology Bombay (IIT Bombay) for providing an Institutional Postdoctoral Fellowship.

References

Adnan et al., 2019 Adnan AI, Ong MY, Nomanbhay S, Chew KW, Show PL. Technologies for biogas upgrading to biomethane: a review. Bioengineering. 2019;6(4):92. Agostini et al., 2020 Agostini A, Giuntoli J, Marelli L, Amaducci S. Flaws in the interpretation phase of bioenergy LCA fuel the debate and mislead policymakers. Int J Life Cycle Assess. 2020;25(1):17–35. Ardolino et al., 2018 Ardolino F, Parrillo F, Arena U. Biowaste-to-biomethane or biowaste-to-energy? An LCA study on anaerobic digestion of organic waste. J Clean Prod. 2018;174:462–476. Arvidsson et al., 2018 Arvidsson R, Tillman AM, Sandén BA, et al. Environmental assessment of emerging technologies: recommendations for prospective LCA. J Ind Ecol. 2018;22(6):1286–1294. Atelge et al., 2020 Atelge MR, Krisa D, Kumar G, et al. Biogas production from organic waste: recent progress and perspectives. Waste Biomass Valoriz. 2020;11(3):1019–1040. Awe et al., 2017 Awe OW, Zhao Y, Nzihou A, Minh DP, Lyczko N. A review of biogas utilisation, purification and upgrading technologies. Waste Biomass Valoriz. 2017;8(2):267–283. Aziz et al., 2019 Aziz NIHA, Hanafiah MM, Gheewala SH. A review on life cycle assessment of biogas production: challenges and future perspectives in Malaysia. Biomass Bioenergy. 2019;122:361–374. Bare, 2002 Bare JC. TRACI: The tool for the reduction and assessment of chemical and other environmental impacts. J Ind Ecol. 2002;6(3-4):49–78. Bauer et al., 2013 Bauer F, Persson T, Hulteberg C, Tamm D. Biogas upgrading– technology overview, comparison and perspectives for the future. Biofuels

Bioprod Bioref. 2013;7(5):499–511. Bedoić et al., 2019 Bedoić R, Čuček L, Ćosić B, et al. Green biomass to biogas– a study on anaerobic digestion of residue grass. J Clean Prod. 2019;213:700– 709. Bond and Templeton, 2011 Bond T, Templeton MR. History and future of domestic biogas plants in the developing world. Energy Sustain Dev. 2011;15(4):347–354. Castellani et al., 2018 Castellani B, Rinaldi S, Bonamente E, Nicolini A, Rossi F, Cotana F. Carbon and energy footprint of the hydrate-based biogas upgrading process integrated with CO2 valorization. Sci Total Environ. 2018;615:404–411. Chen et al., 2015 Chen XY, Vinh-Thang H, Ramirez AA, Rodrigue D, Kaliaguine S. Membrane gas separation technologies for biogas upgrading. RSC Adv. 2015;5(31):24399–24448. Cherubini, 2010 Cherubini F. GHG balances of bioenergy systems–overview of key steps in the production chain and methodological concerns. Renew Energy. 2010;35(7):1565–1573. Collet et al., 2017 Collet P, Flottes E, Favre A, et al. Techno-economic and Life Cycle Assessment of methane production via biogas upgrading and power to gas technology. Appl Energy. 2017;192:282–295. Costa et al., 2018 Costa MP, Schoeneboom JC, Oliveira SA, Vinas RS, de Medeiros GA. A socio-eco-efficiency analysis of integrated and non-integrated crop-livestock-forestry systems in the Brazilian Cerrado based on LCA. J Clean Prod. 2018;171:1460–1471. Dębowski et al., 2013 Dębowski M, Zieliński M, Grala A, Dudek M. Algae biomass as an alternative substrate in biogas production technologies. Renew Sustain Energy Rev. 2013;27:596–604. Ekvall, 2020 Ekvall, T., 2020. Attributional and consequential life cycle assessment. Sustainability Assessment at the 21st century, 395, pp. 1–22. European Commission, 2010 European Commission, 2010. t research centre–institute for environment and sustainability. ILCD handbook.

International Reference Life Cycle Data System. Framework and Requirements for Life Cycle Impact Assessment Models and Indicators. Ferreira et al., 2019 Ferreira AF, Toledo-Cervantes A, de Godos I, Gouveia L, Munõz R. Life cycle assessment of pilot and real scale photosynthetic biogas upgrading units. Algal Res. 2019;44:101668. Florio et al., 2019 Florio C, Fiorentino G, Corcelli F, et al. A life cycle assessment of biomethane production from waste feedstock through different upgrading technologies. Energies. 2019;12(4):718. Grande, 2012 Grande CA. Advances in pressure swing adsorption for gas separation. ISRN Chem Eng. 2012;2012:982934. Groen and Heijungs, 2017 Groen EA, Heijungs R. Ignoring correlation in uncertainty and sensitivity analysis in life cycle assessment: what is the risk?. Environ Impact Assess Rev. 2017;62:98–109. Guinée, 2001 Guinée J. Handbook on life cycle assessment–operational guide to the ISO standards. Int J Life Cycle Assess. 2001;6(5):255. Gunnarsson et al., 2014 Gunnarsson IB, Alvarado-Morales M, Angelidaki I. Utilization of CO2 fixating bacterium Actinobacillus succinogenes 130Z for simultaneous biogas upgrading and biosuccinic acid production. Environ Sci Technol. 2014;48(20):12464–12468. Hijazi et al., 2016 Hijazi O, Munro S, Zerhusen B, Effenberger M. Review of life cycle assessment for biogas production in Europe. Renew Sustain Energy Rev. 2016;54:1291–1300. ISO, 2006 ISO (2006). ISO 14044—environmental management—life cycle assessment—requirements and guidelines. International Standard Organization. Jonsson and Westman, 2011 Jonsson, S., Westman, J., 2011. Cryogenic Biogas Upgrading Using Plate Heat Exchangers (Master’s thesis). Chalmers University of Technology, Göteborg. Kadam and Panwar, 2017 Kadam R, Panwar NL. Recent advancement in biogas enrichment and its applications. Renew Sustain Energy Rev. 2017;73:892–903.

Khan et al., 2017 Khan IU, Othman MHD, Hashim H, et al. Biogas as a renewable energy fuel–a review of biogas upgrading, utilisation and storage. Energy Convers Manage. 2017;150:277–294. Lelek and Kulczycka, 2020 Lelek L, Kulczycka J. Life cycle modelling of the impact of coal quality on emissions from energy generation. Energies. 2020;13(6):1515. Leonzio, 2016 Leonzio G. Upgrading of biogas to bio-methane with chemical absorption process: simulation and environmental impact. J Clean Prod. 2016;131:364–375. Liang et al., 2019 Liang L, Wang Y, Ridoutt BG, et al. Agricultural subsidies assessment of cropping system from environmental and economic perspectives in North China based on LCA. Ecol Indic. 2019;96:351–360. Lombardi and Francini, 2020 Lombardi L, Francini G. Techno-economic and environmental assessment of the main biogas upgrading technologies. Renew Energy. 2020;156:440–458. Lorenzi et al., 2019 Lorenzi G, Gorgoroni M, Silva C, Santarelli M. Life cycle assessment of biogas upgrading routes. Energy Proc. 2019;158:2012–2018. Luo et al., 2012 Luo G, Johansson S, Boe K, Xie L, Zhou Q, Angelidaki I. Simultaneous hydrogen utilization and in situ biogas upgrading in an anaerobic reactor. Biotechnol Bioeng. 2012;109(4):1088–1094. Lyng and Brekke, 2019 Lyng KA, Brekke A. Environmental life cycle assessment of biogas as a fuel for transport compared with alternative fuels. Energies. 2019;12(3):532. Marconi and Favi, 2020 Marconi M, Favi C. Eco-design teaching initiative within a manufacturing company based on LCA analysis of company product portfolio. J Clean Prod. 2020;242:118424. Masebinu et al., 2014 Masebinu SO, Aboyade A, Muzenda E. Enrichment of biogas for use as vehicular fuel: a review of the upgrading techniques. Int J Res Chem Metall Civ Eng. 2014;1(1):88–97. Meegoda et al., 2018 Meegoda JN, Li B, Patel K, Wang LB. A review of the

processes, parameters, and optimization of anaerobic digestion. Int J Environ Res Public Health. 2018;15(10):2224. Miltner et al., 2017 Miltner M, Makaruk A, Harasek M. Review on available biogas upgrading technologies and innovations towards advanced solutions. J Clean Prod. 2017;161:1329–1337. Molino et al., 2013 Molino A, Nanna F, Ding Y, Bikson B, Braccio G. Biomethane production by anaerobic digestion of organic waste. Fuel. 2013;103:1003–1009. Møller et al., 2009 Møller J, Boldrin A, Christensen TH. Anaerobic digestion and digestate use: ing of greenhouse gases and global warming contribution. Waste Manage Res. 2009;27(8):813–824. Montgomery and Bochmann, 2014 Montgomery LF, Bochmann G. PreTreatment of Feedstock for Enhanced Biogas Production Ireland: IEA Bioenergy; 2014;1–20. Navas-Anguita et al., 2019 Navas-Anguita Z, Cruz PL, Martin-Gamboa M, Iribarren D, Dufour J. Simulation and life cycle assessment of synthetic fuels produced via biogas dry reforming and Fischer-Tropsch synthesis. Fuel. 2019;235:1492–1500. Nie et al., 2013 Nie H, Jiang H, Chong D, Wu Q, Xu C, Zhou H. Comparison of water scrubbing and propylene carbonate absorption for biogas upgrading process. Energy Fuels. 2013;27(6):3239–3245. Okeke and Mani, 2017 Okeke IJ, Mani S. Techno-economic assessment of biogas to liquid fuels conversion technology via Fischer-Tropsch synthesis. Biofuels Bioprod Bioref. 2017;11(3):472–487. Persson et al., 2006 Persson M, Jönsson O, Wellinger A. Biogas Upgrading to Vehicle Fuel Standards and Grid Injection. Vol. 37 IEA Bioenergy 2006;1–34 December. Petersson and Wellinger, 2009 Petersson A, Wellinger A. Biogas upgrading technologies–developments and innovations. IEA Bioenergy. 2009;20:1–19. Petrillo et al., 2016 Petrillo A, De Felice F, Jannelli E, Autorino C, Minutillo M,

Lavadera AL. Life cycle assessment (LCA) and life cycle cost (LCC) analysis model for a stand-alone hybrid renewable energy system. Renew Energy. 2016;95:337–355. Pizzol et al., 2017 Pizzol M, Laurent A, Sala S, Weidema B, Verones F, Koffler C. Normalisation and weighting in life cycle assessment: quo vadis?. Int J Life Cycle Assess. 2017;22(6):853–866. Rajendran et al., 2020 Rajendran K, Mahapatra D, Venkatraman AV, Muthuswamy S, Pugazhendhi A. Advancing anaerobic digestion through twostage processes: current developments and future trends. Renew Sustain Energy Rev. 2020;123:109746. Reap et al., 2008 Reap J, Roman F, Duncan S, Bras B. A survey of unresolved problems in life cycle assessment-Part 1: goal and scope and inventory analysis. Int J Life Cycle Assess. 2008;13(4):290–300. Ryckebosch et al., 2011 Ryckebosch E, Drouillon M, Vervaeren H. Techniques for transformation of biogas to biomethane. Biomass Bioenergy. 2011;35(5):1633–1645. Sahota et al., 2018 Sahota S, Shah G, Ghosh P, et al. Review of trends in biogas upgradation technologies and future perspectives. Bioresour Technol Rep. 2018;1:79–88. Scholwin and Held, 2013 Scholwin F, Held J. Biomethane from anaerobic processes. In: Kaltschmitt M, Themelis NJ, Bronicki LY, Söder L, Vega LA, eds. Renewable Energy Systems. New York: Springer; 2013. Scholz et al., 2013 Scholz M, Melin T, Wessling M. Transforming biogas into biomethane using membrane technology. Renew Sustain Energy Rev. 2013;17:199–212. Seabra et al., 2011 Seabra JE, Macedo IC, Chum HL, Faroni CE, Sarto CA. Life cycle assessment of Brazilian sugarcane products: GHG emissions and energy use. Biofuels Bioprod Bioref. 2011;5(5):519–532. SoCalGas, 2020 SoCalGas. Power-to-Gas Technology. 2020. https://www.socalgas.com/smart-energy/renewable-gas/power-to-gas (accessed 20.07.20).

Sridhar et al., 2007 Sridhar S, Smitha B, Aminabhavi TM. Separation of carbon dioxide from natural gas mixtures through polymeric membranes—a review Separation. Purif Rev. 2007;36(2):113–174. Starr et al., 2012 Starr K, Gabarrell X, Villalba G, Talens L, Lombardi L. Life cycle assessment of biogas upgrading technologies. Waste Manage. 2012;32(5):991–999. Tambone et al., 2010 Tambone F, Scaglia B, D’Imporzano G, et al. Assessing amendment and fertilizing properties of digestates from anaerobic digestion through a comparative study with digested sludge and compost. Chemosphere. 2010;81(5):577–583. Uggetti et al., 2016 Uggetti E, os F, Solé M, García J, Ferrer I. Biogas from algae via anaerobic digestion. Algae Biotechnology Cham: Springer; 2016;195– 216. Walsh et al., 2012 Walsh JJ, Jones DL, Edwards-Jones G, Williams AP. Replacing inorganic fertilizer with anaerobic digestate may maintain agricultural productivity at less environmental cost. J Plant Nutr Soil Sci. 2012;175(6):840– 845. Wang et al., 2016 Wang QL, Li W, Gao X, Li SJ. Life cycle assessment on biogas production from straw and its sensitivity analysis. Bioresour Technol. 2016;201:208–214. Wang et al., 2020 Wang C, Malik A, Wang Y, et al. The social, economic, and environmental implications of biomass ethanol production in China: a multiregional input-output-based hybrid LCA model. J Clean Prod. 2020;249:119326. Wellinger and Murphy, 2013 Wellinger A, Murphy. J.D. In: Baxter D, ed. The biogas handbook: science, production and applications. Elsevier 2013. Wiley et al., 2011 Wiley PE, Campbell JE, McKuin B. Production of biodiesel and biogas from algae: a review of process train options. Water Environ Res. 2011;83(4):326–338. Wu et al., 2017 Wu M, Zhang W, Ji Y, et al. Coupled CO2 fixation from ethylene oxide off-gas with bio-based succinic acid production by engineered recombinant Escherichia coli. Biochem Eng J. 2017;117:1–6.

Wulf et al., 2018 Wulf C, Linssen J, Zapp P. Power-to-gas—concepts, demonstration, and prospects. Hydrogen Supply Chains Academic Press 2018;309–345. Zhang et al., 2017 Zhang X, Bauer C, Mutel CL, Volkart K. Life cycle assessment of power-to-gas: approaches, system variations and their environmental implications. Appl Energy. 2017;190:326–338. Zhang et al., 2020 Zhang B, Su S, Zhu Y, Li X. An LCA-based environmental impact assessment model for regulatory planning. Environ Impact Assess Rev. 2020;83:106406.

Further reading

Feiz and Ammenberg, 2017 Feiz R, Ammenberg J. Assessment of feedstocks for biogas production, part I—a multi-criteria approach. Resources Conserv Recycl. 2017;122:373–387. Herrmann et al., 2016 Herrmann C, Idler C, Heiermann M. Biogas crops grown in energy crop rotations: linking chemical composition and methane production characteristics. Bioresour Technol. 2016;206:23–35. Rafique et al., 2010 Rafique R, Poulsen TG, Nizami AS, Murphy JD, Kiely G. Effect of thermal, chemical and thermo-chemical pre-treatments to enhance methane production. Energy. 2010;35(12):4556–4561. Scarlat et al., 2018 Scarlat N, Dallemand JF, Fahl F. Biogas: developments and perspectives in Europe. Renew Energy. 2018;129:457–472.

Chapter 16

The role of techno-economic implications and governmental policies in accelerating the promotion of biomethane technologies

Dhamodharan Kondusamy¹, ², Mehak Kaushal³, Saumya Ahlawat⁴ and Karthik Rajendran⁵, ¹1Department of Civil Engineering, Indian Institute of Technology Guwahati, Guwahati, India, ²2Institute of Soil, Water and Environmental Science, Agricultural Research Organization, Israel, ³3System Biology for Biofuel Group, International Centre for Genetic Engineering and Biotechnology, New Delhi, India, ⁴4Department of Biosciences and Bioengineering, Indian Institute of Technology Guwahati, Guwahati, India, ⁵5Department of Environmental Science, SRM UniversityAP, Mangalagiri, India

Abstract

Biomethane production and upgrading technologies have gained significance in Europe, since the introduction of the EU landfill directive in 1999. This directive banned the landfilling of organic wastes in the EU region; failing to comply with the directive incurred a penalty of €75/t. Most renewable energy technologies face hurdles in competing with conventional technologies. From an industrial perspective for commercialization, the technology needs to be economically viable. If it is not economically viable, the industry needs start-up through policies and incentives. Biomethane is no different from other renewable energy systems where initial policy is needed to kick-start the industry in any country. This chapter deals with the role of techno-economic studies and effective policies in providing a decision- system for successful implementation of biomethane technologies.

Keywords

Techno-economic analysis; biomethane; policy ; energy policy; biomethane upgrading

Chapter outline

Outline


16.2 Role of techno-economic studies in anaerobic digestion 449

16.2.1 Feedstocks 450 16.2.2 Gas purification technology 450 16.2.3 Biogas utilization 451 16.2.4 Subsidies 452

16.3 Successful policies in anaerobic digestion implementation 452

16.3.1 Policies and regulations 452 16.3.2 Renewable energy-related policies and regulations 453

16.3.3 Agriculture policies and regulations 454 16.3.4 Waste management policies 454 16.3.5 Incentives 455 16.3.6 Policy instruments introduced in various countries as a to AD industry growth 456

16.4 Decision- system for biomethane implantation with technoeconomic analysis and policies 461

16.5 Conclusion 462

References 463

16.1 Introduction

By 2020, every European Union state had a target to ensure that 20% and 10% of their total energy utilization and transportation requirements, respectively, were fulfilled by bioenergy. Biomethane is one of the bioenergy sources that contribute toward this mandate. By 2030 this share will increase to 32% and 14% of total energy consumption and transport sector consumption, respectively (2018/844/EC, REDII). For a technology to be commercialized different aspects are analyzed including technical, environmental, market, financial, and legal aspects. Financial feasibility depends on the capital, operational, and maintenance costs, which are the important factors governing the suitability of a project. The amount of biomethane generated depends on biogas composition, which in turn depends upon the raw material and production process used. Methane amount varies from 45%–60% and 60%–70% in biogas from landfills and organic residue-based digesters, respectively (Khan et al., 2017). In the light of current consumption trends, natural gas faces the fastest increase (50%) in demand among fossil fuels by 2040 (IEA, 2009). Hence, it is imperative that all countries work toward a carbon-neutral transport sector and diversify energy supply. In Europe, 19% of natural gas produced is utilized for transportation purposes (Pellegrino et al., 2017). To that end, governmental policies and management practices should be designed to encourage green energy generation. Biogas-biomethane provides a carbon-negative substitute to natural gas as it reduces greenhouse gases in amounts equivalent to 200 g of CO2eq/kWh generated (200 gCO2eq/kWh) (Adelt et al., 2011). This biomethane is either injected into the gas grid network or compressed to be used as transport fuel. In comparison to diesel and petrol, compressed natural gas when used as transport fuel reduces GHG emissions by 21%–24% (32 gCO2/kWh to 36 gCO2/kWh). When used as vehicle fuel, a 20% biomethane mixture decreases emissions by 24 gCO2/kWh in comparison to fossil fuels. Similarly, 100% biomethane usage can cause emissions reductions of up to 119 gCO2/kWh (DENA, 2011). Biogas upgrading to biomethane is more advantageous than using biogas directly in heat and power units as it enables a reduction of GHG, NOx, and particulate matter emission and thus is more environmentally friendly (Ravina and Genon,

2015). Prevalent technologies for biomethanation, which are currently employed in the European region, include water scrubbing (40%), pressure swing adsorption (PSA) (25%), chemical scrubbing (25%), physical scrubbing (6%), and membrane separation (4%) (Niesner et al., 2013). In spite of high methane losses of around 3%–10% volume, the most common method of biogas upgradation in Sweden is PSA technology. Use of water scrubbing for upgrading has the advantage of high methane purity (>99 vol.%), however, it has the disadvantages of high wastewater generation and power requirement. Similarly, amine scrubbing provides pure methane but has similar disadvantages of high power requirements and use of organic solvents. For carbon footprint lifecycle assessment, a process is rendered unsustainable if it leaks more than 4% methane into the atmosphere through insufficient transport methods or power generation methods, or during the purification process (Ravina and Genon, 2015). This is because of the high global warming potential of methane, which negatively affects the environmental and economic feasibility. As a means of surplus electricity storage in the form of methane, direct methanation of biogas was found to be more economical in comparison to carbon dioxide capture and its subsequent methanation (Vo et al., 2018). Membrane processes are considered suitable for biogas generation units with less than 1000 m³(STP)/h capacity (Miltner et al., 2017). Carbon membranes, when used for biogas upgradation at 8.5 bar feed pressure, incur a low processing cost of $0.078/m³. Biomethane, upgraded from biogas, can be used as a platform fuel because it can undergo conversion to other fuels through steam reforming and catalytic processing (Ferella et al., 2017). Vo et al. (2018) conducted process simulation with different feedstocks and different biomethanation methods and found that, among the different scenarios tested, the biogas plant using grass and dairy slurry as biomass source and amine scrubber for upgrading produced the cheapest biomethane (€0.76/m³) compared to amine scrubbing with CO2 directed to ex situ biological methanation and biogas upgrading through ex situ biological methanation. This chapter highlights the various roles of techno-economic assessment and policies prevailing across the world for promoting biomethane technologies. To predict production costs and understand the economic and technical limits of second-generation biofuels (i.e., agricultural and organic waste), many mathematical approaches have been used to obtain supply–cost curves. The outcome of a technical limit study was then incorporated into a geographical information system to obtain the energy distribution. There are several components involved in completing a technology, where “one size does not fit

all.” The cost and policies prevailing in one country might not be implemented in other countries due to several constraints. However, the structure and mechanisms of economics and policies mentioned in this chapter should help to provide an understanding of the existing methods followed in various countries.

16.2 Role of techno-economic studies in anaerobic digestion

Economic growth has aided the emergence of diverse enterprises and sectors, which are aimed toward achieving their goals with minimal operational costs. One of the means to reduce cost is energy independence, that is, decentralized electricity generation, which reduces operational costs. Several authors have conducted techno-economic analyses in relation to this project. Different components of techno-economic analysis, such as feedstock, gas purification technology, and biogas utilization, are described in detail below.

16.2.1 Feedstocks

The depletion of conventional energy sources and increase in greenhouse gas emissions have spurred the development of sustainable green energy sources. One of these sources is biomass-based fuels as a renewable source of energy. Globally, the estimation of energy from biomass produced annually is almost eight times the annual energy requirement. By 2050, this stored energy will be able to fulfill 10%–20% of the global energy demand (Akay et al., 2005). This has promoted the use of agro-based and agro-industrial biomass waste for biogas generation via anaerobic digestion; biogas can further be used for producing electricity, heat, and biomethane (Scarlat et al., 2015). Use of agro-waste, municipal waste, animal waste, and manure residue for energy production serves an additional purpose: that of waste treatment (Fan et al., 2018). Biogas can also be generated using biomass feedstocks like sugarcane, corn, sugar beet, etc., however, this raises the food-versus-fuel debate (Bušić et al., 2018). A number of authors have undertaken economic and sensitivity analyses for biomethane production. Murphy and Power (2008) studied the optimization of biogas production using three different crop rotations and conducted sensitivity analysis with respect to varying digestate cost and concluded that biomethane production

from the plantation of wheat, barley, and sugar beet in rotation for the cultivation of 48,000 ha resulting in 17% more transport fuel in Ireland. Georgakakis et al. (2003) developed a net present value (NPV)-based model for financial analysis to study the economics of a pig manure-fed biogas plant. According to this study the feedstock and digestate logistics are important parameters when analyzing the financial, environmental, and social feasibility of a plant. Poliafico and Murphy (2007) reported a maximum distance of 15–25 km to be feasible as a higher transport distance incurs higher fuel costs and adds to GHG emissions. From the techno-economic analysis of feedstocks, different feedstocks have different biomethane -producing capacities and the logistics of feedstocks costs more, and so if the biomethanation plant is close to the source of feedstock it would be beneficial financially.

16.2.2 Gas purification technology

Technologies available for biomethanation at large scale include adsorption (physical and chemical), membrane separation, and cryogenic process. While high-pressure water scrubbing is the most widely used physical adsorption method, amine scrubbing is most popular among the chemical adsorption processes. PSA is an adsorption method using sorbent materials such as activated carbon, silica gel, alumina, resins, and zeolites (Ferella et al., 2017). Carbon dioxide separation from methane can also be carried out using different membrane types such as polymeric, ceramic, inorganic, cross-linked, metallicbased, and carbon-based. Another method of carbon dioxide separation is the cryogenic process that takes advantage of the difference between cryogenic temperatures of carbon dioxide and methane. Lee (2017) reported biomethane to be a better energy source in comparison to biogas for the following reasons: higher biomethane retail price, zero purification costs, and higher profit generation compared to biobutanol, biohydrogen, and biodiesel. Upgrading units handling great/vast quantities of biogas require a threshold quantity of natural gas to drive a supplementary boiler that generates thermal power for the thermophilic anaerobic digestion process. The electrical energy requirement of a biomethane production plant is proportional to the size of the upgrading unit and hence large plants incur more energy costs. Financial analysis revealed that the

biomethane price governs the optimum ratio of upgrading the unit size and total production of the plant. The upgrading unit should be kept to a minimum size (<250 m³/d), if the biomethane retail price is lower than $0.45/Nm³ (Baccioli et al., 2019). The profitability of biomethane is feasible with a retail price range of $1.3–$1.96/Nm³ along with a provision for a feed-in tariff which provides a carbon credit of €99/ton of carbon dioxide. From the studies it is clear that water scrubbing and PSA technology occupies a major portion of biomethane projects due to the cost and efficiency.

16.2.3 Biogas utilization

Biogas production takes places by microbial cultivation under anaerobic conditions. The raw material for microbial growth consists of organic matter such as animal waste, food waste, and wastewater. Biogas composition varies depending upon the raw material used, but generally consists of methane (50%– 70%), carbon dioxide (25%–50%), hydrogen (1%–5%), nitrogen (1%–5%), and hydrogen sulfide (0.3%–3%) in volume basis. Biogas power plant costs include capital investment and operating costs. Capital investment includes the cost of the engine generator and installation, while operating costs include fixed and variable costs. Fixed operating costs include maintenance costs, depreciation cost, and overhead expenses, and varies yearly because of depreciation. The operating cost of power plants is influenced by the fuel requirements and is calculated from the diesel to biogas blending ratio per kWh. Rotund et al. (2017) studied a biogas plant of 120 m³/h capacity with water scrubbing technology for purification and assessed the cost based on different end uses of biomethane. It was found that biomethane production cost per unit was lower when used for grid injection (0.54 V/m³) in comparison to its end use as a transportation fuel (0.73 V/m³). Cucchiella et al. (2018) undertook a profitability analysis of a 150 m³/h-capacity biomethane plant with the end application being injecting the gas to the grid. The study found that plants using manure residues, solely or in combination with energy crops, were unprofitable. Plant size was found to affect NPV, while subsidies were reported to be significant for economic analysis. It was emphasized that financial in the form of subsidies was imperative to ensure economic competitiveness and market penetration of biomethane

production (Fubara et al., 2018).

16.2.4 Subsidies

Use of biomethane as a transportation fuel is hindered by economic feasibility. Different instruments, which can be employed to favor economic feasibility, include lower taxation, subsidies, reduced external costs, diversion of waste biomass feedstocks toward biomethane production, etc. (Uusitalo et al., 2015). Studies conducted to assess the profitability of biomethane plants under different conditions have shown that it is strongly linked with subsidies. In the case of NPV-positive plants, total biomethane generated was found to be positively correlated with incentive level, plant size, and volatile solid ration, while it was negatively correlated with silage price. Shea et al. (2017) found that at despite using different combinations of silage price, plant size, or volatile solid ratio, the biomethane plant was unable to achieve a positive NPV if the incentive was €20/MWh in Ireland. In 1988, Denmark set up Biogas Action Program and became the first country to focus on organized application of anaerobic digestion for agro-wastes. The program focused on research and development, construction and monitoring biogas plants, and spreading awareness about anaerobic digestion. Economic included endowments for setting up biogas plants (up to 40% of costs), profitable loan plans, tax waivers, and subsidies. The EU does not prescribe maintenance of agro-wastes in accordance with nutrient management policies. Nevertheless, subsidies for anaerobic digestion have been prompted by many countries as a means for agro-waste management. The Government of India has provided economic relief to small-scale biogas plants for cooking purposes in the form of subsidies under the National Biogas and Manure Management Program, as shown in Table 16.1.

Table 16.1

S. no.

Central financial assistance (CFA) and states/regions and categories

A.

Central subsidy rates applicable

1.

Northeastern (NE) region states, Sikkim (except plain areas of Assam) and including scheduled caste (SC) and

2.

Plain areas of Assam

3.

Jammu & Kashmir, Himachal Pradesh, Uttarakhand, Nilgiri of Tamil Nadu, Sadar Kurseong and Kalimpong S

4.

Scheduled castes/Scheduled Tribes of all other States except NE Region States (including Sikkim)

5.

All others

B.

Turn-key job fee including warranty for 5 years and quality control (in Rs. per plant)

C.

Additional subsidy (CFA) for toilet-linked biogas plants (in Rs. per plant)

16.3 Successful policies in anaerobic digestion implementation

16.3.1 Policies and regulations

The successful development of the biogas industry in most developed countries has been possible due to a collaborated effort of strategic planning and policy enforcement by the government sector.

16.3.2 Renewable energy-related policies and regulations

16.3.2.1 Renewable energy generation targets

Since biogas serves as an alternative to natural gas both as a complete replacement or in a mixture, it is used for heat/electricity generation as well as being a transportation fuel. Various policies and regulations have been adopted by countries around the world to encourage the use of energy/fuel derived from renewable sources and in turn this has stimulated the expansion of AD technology. Specific renewable energy targets (RETs) have been set by different countries wherein they claim to increase their share of the renewable energy within a certain period of time. For instance, the RETs to be achieved by 2020 for the European Union and the United States are 17% and 18%, respectively, of the total energy used (Vasco-Correa et al., 2018). Brazil and New Zealand aim to

reach targets of 70% and 90%, respectively, for renewable energy by 2020 (Vasco-Correa et al., 2018).

16.3.2.2 Greenhouse gas emission reduction targets

According to the 2012 Doha amendment to the Kyoto Protocol, 37 countries have binding targets to reduce their share of greenhouse gases (GHGs) by 2020. This has also promoted the use of AD due to its environmentally friendly GHG saving potential (European Parliament, 2009). , the United Kingdom, and Finland have vowed to decrease their GHG to volumes 80% below their 1990 levels by 2050, while the United States has a target of a 17% reduction in GHG emissions to be reached by 2020 (Vasco-Correa et al., 2018).

16.3.2.3 Rural development

Rural areas have an abundance of decomposable feedstocks which serve as potential substrates for AD technology. Thus, countries provide financial assistance in the form of grants, loans, and training to farmers for establishing and running AD facilities in rural areas (Vasco-Correa et al., 2018).

16.3.3 Agriculture policies and regulations

The agriculture and livestock sector results in the creation of pollutants which, if discharged to the land, water bodies, or atmosphere, could result in serious environmental damage and health complications. Thus, countries have strict air and water emission policies which promote the adoption of best management

practices for feedstock collection, handling and storing of manure, and methods to decrease odors, liquid waste, and dust, for instance, the Nitrate Directives 1991 regulation existing in Europe and the Clean Water Act of the United States (Vasco-Correa et al., 2018). Some countries have restrictions on the nutrient level of animal manure, for instance, setting limits on nitrate and phosphate fertilizer usage. These types of legislation place an onus on agriculture to adopt the best management practices, including use of AD for agrobusiness and livestock management, thus minimizing air, water, and soil pollution.

16.3.4 Waste management policies

Policies and regulations that restrict the disposal of organic materials such as municipal solid waste to landfills also play an important role in driving the development of AD treatment of these materials (Vasco-Correa et al., 2018). The increase in landfill levies in some countries has a direct correlation with the number of AD plants established (Edwards et al., 2015). Certain cities have banned organic waste from entering landfill before undergoing treatment at a compost or AD facility.

16.3.5 Incentives

For the AD industry to sustain and make a profit it is a prerequisite for the government of various countries to provide in the form of incentive programs as described below.

16.3.5.1 Feed-in tariff (FIT)

FITs are policy instruments intended at accelerating the investment in renewable energy technologies. Under these incentive schemes contracts are offered to renewable energy producers based on the total cost of production of the technology (Vasco-Correa et al., 2018). For instance, a lower per-kWh price is given for AD over other technologies such as wind energy, landfill gas, and hydropower.

16.3.5.2 Credits for carbon reduction and carbon trading

Giving vendible carbon credits to producers who use technologies that limit the emission of GHGs or generate electricity using renewable sources is an important step toward driving the adoption of AD at various agro-industrial facilities or large farms.

16.3.5.3 Tax exemptions and tax credits

Tax exemptions and tax credits for businesses that are producers or consumers of renewable energy have boosted various renewable technologies including AD.

16.3.5.4 Credits for renewable energy and renewable transportation fuel

The adoption of AD is further monetized and promoted by offering credits for use of biogas as a renewable transportation fuel, in turn resulting in a reduction

of fossil fuel usage.

16.3.5.5 Credits for nutrient load reduction

Credits are also provided to businesses/industries for reducing the nutrient load of the effluent that is discharged to water bodies. The producers adopt AD technology to digest the organic matter present in manure or waste, serving a dual purpose of reduced nutrient content and electricity generation.

16.3.5.6 Renewable heat incentive

Some countries also provide a renewable heat incentive (RHI) to owners of buildings who use renewable energy sources for heat generation in place of using fossil fuels or electricity from the grid (UK DECC, 2015).

16.3.6 Policy instruments introduced in various countries as a to AD industry growth

Europe is the world leader in biogas production, followed by Asia and America. The various regulations and policies adopted by various countries are discussed below.

16.3.6.1

is the top producer of biogas in the world. This has been possible through various innovative policies and incentives offered based on plant size, raw material usage, and technology employed to the biogas producers and consumers. Through the Biofuels Quota Act of 2007, has a mandate of a minimum biofuel share to be sold in the market (Capodaglio et al., 2016). The Renewable Energy Sources Act of 2000 (RESA) offers incentives such as FIT and priority connection rights for electricity generated from renewable sources. also offers a higher FIT rate for AD-derived electricity than that derived from wind or hydropower (Edwards et al., 2015). After the implementation of the RESA act in 2000, every year from 2001 to 2004, 250 new AD plants were constructed. In 2004, the RESA act was amended to give different FITs according to the electricity generation capacity (150 kW, 500 kW, 5000 kW, and 20,000 kW) of the plant. This revision of the act led to 450 new plants being established every year from 2004 to 2009. Further, the FIT was revised on regular intervals for small plants to encourage farmers to adopt the technology (IEA, 2009; IEA, 2012). Meanwhile for large-scale plants having a capacity of more than 20,000 kW the FIT was gradually reduced from 2004 and was completely stopped in 2017. This was done to boost both technology as well as balance the competition between small- and large-scale AD producers. Bonuses were given on the usage of selected raw material such as animal manure, plant biomass, and energy crops. Bonuses were also given depending on the kind of technology used and for use of dual heat and power facilities. These bonuses were mostly valid for plant sizes less than 20,000 kW, thus ensuring new facilities are being built and centralization of production is restricted. These measures resulted in further growth of the number of new plants being established to 1000 per year from 2009 to 2012 (Fachverband Biogas, 2017). After 2015, removed the incentives for upgrading of the large plants, along with a 1% reduction in FIT every year after 2012. This resulted in a decline in the number of new plants however, due to the demand for technology, the biogas market was able to sustain itself. biogas policy has been successful in generating revenue through production and innovation and has also been a guiding light for other countries through technology transfer and consultancy services. In 2010, “Initiative for Natural-Gas-Based Mobility” was implemented by to promote the market share of natural gas vehicles. This was directed toward an increase in the methane production to be used in transportation fuel with and without blending (Le Fevre, 2014).

16.3.6.2 The United States

The United States has maintained the position of being the second largest biogas producer in the world for a long time. The AD industry in the United States is well established in utilizing sewage waste as substrate and coupling heat and electricity facilities together (Edwards et al., 2015). The development of AD in this country is driven by two key regulations. One is the renewable portfolio standard (RPS) through which FITs are provided and another through which renewable energy mandates are set. In 2014, as a strategy to fight climate change by reduction of GHG emissions by 25% by 2020, the United States took fasttrack measures for the adoption of the AD industry (USDA, 2014). The United States does not have strict agro-environmental rules like the EU, however it does have regulations restricting surface and ground water pollution, such as the Clean Water Act and its follow-up amendments. The major growth of the AD sector (AD digesters increased from less than 30 in 2003 to 191 in 2012) in the United States took place after 2000, when the key legislations were employed to the development of new plants through financial assistance in the form of loans, grants, tax rebates, and carbon credits (Bracmort, 2010). One of the significant national policies for driving AD is the Renewable Electricity Production Tax Credit (PTC) through which tax credit is given for a period of 10 years to new plants (Bangalore et al., 2016). Apart from federal polices, most states have their own state policy for AD development. States have their own mandates for renewable energy, for instance, Pennsylvania and California have goals to produce 18% and 33% of their energy, respectively, from renewable sources by 2020. Similarly, different states have set different targets for GHG emissions reductions as well. The remuneration rates for a 25-year contract vary from 9.3 cents/kWh in California to 14 cents/kWh in Vermont (Bangalore et al., 2016). Wisconsin has the highest number of AD digesters in the United States due to its generous tax credits, incentives, and moderate targets for renewable energy. Regardless of these policy instruments, the future for AD is uncertain, due to the low price of abundantly available natural gas.

16.3.6.3 The United Kingdom

The Office of Gas and Electricity Markets (OFGEM) in the United Kingdom manages a renewable funding scheme through the Energy Act 2008 (Hermann and Hermann, 2018). The AD industry in this country gets its monetary through incentives for electricity [Renewables Obligation (RO) and the FIT] and heat production from biogas (RHI). The FIT scheme was introduced in 2010 and was mainly aimed at extending to small-scale energy-generating plants, while the RO system along with the tax regulation mechanism was applied for the larger setups resulting in more than 5 MW of energy (Hermann and Hermann, 2018). New incentives are being introduced in order to upgrade biogas plants to biomethane production which is also a viable step looking at the widespread gas distribution network in the United Kingdom. In 2017, the Clean Growth Strategy was announced by the UK government under which £2.5 billion would be directed toward cleaner, smarter, and flexible power, a move aiming at reducing the dependence on a high carbon-emitting transport system (Damave, 2018). To promote the gas-to-grid industry in the United Kingdom, higher incentives are given for injection of biomethane to grid rather than direct heat (OGEM, 2017). Further, the United Kingdom has imposed restrictions on the quantity of nitrates that can be diverted to surface or groundwater through manure disposal. Although this has put a limitation on the time for which manure can be stored, farmers get easy excess to transfer manure to neighboring sites without the license prerequisite (Edwards et al., 2015; Vasco-Correa et al., 2018). The United Kingdom also has strict landfill regulations, with heavy landfill levies which have boosted the diversion of municipal/agricultural waste from landfill to energy-generating AD plants (Edwards et al., 2015).

16.3.6.4 Italy

The energy policy in Italy was revised in 2013 to plants on the basis of size and the raw material used. Small plants and usage of manure were promoted through plant- and feedstock-based incentives and subsidies. Three areas of

manure usage were adopted: below 50% manure usage, above 50% manure usage, and complete 100% manure usage. With increasing manure usage incentives were increased, while the same decreased with increasing plant size. For maximum profits of AD plants, the reuse of biogas slurry and increased manure recovery rate were encouraged (Bartoli et al., 2019). Further, a tradable certificate scheme (CIC) is also offered for biofuel producers where certificates are granted on the basis of manure share ranging from below 70% to 100%. Biomethane producers directly distributing the gas rather than taking up the role of wholesalers are offered an increased (50%) certificate value for a period of 10 years, while for new plants additional incentives and certificates are granted. The 2013 policy was revised in 2018, raising the incentives for annual biomethane production (MISE, 2018). The key highlights of the new decree are a €375 payment per CIC for a period of 10 years, double subsidy for the use of certain substrates (i.e., organic fraction of municipal solid waste), and a supplementary incentives for producers-cum-distributors of biomethane (MISE, 2018).

16.3.6.5 Sweden

Biogas systems are well established in Sweden, where more than half of the biogas is used as transportation fuel (Larsson et al., 2016). This has been possible largely due to cooperative measures at both government and municipality levels. Regional initiatives include good management of substrate collection along with biogas production and distribution. National waste legislation restricts the diversion of waste to landfills and has set a minimum share of organic waste to be treated biologically (Swedish Environmental Protection Agency, 2017). Regional authorities have been actively promoting the AD industry as a solution to transport problems and this has resulted in large biogas production plants being established in recent years (Larsson et al., 2016).

16.3.6.6 China

It was after the 1970s that the biogas industry grew in China at a rapid pace and biogas digesters were installed in rural areas (Gu et al., 2016). This was ed by government measures aimed at overcoming the energy scarcity problems faced in these areas. The development of biogas infrastructure has been made a part of the long-term national developmental schemes by the government (Jiang et al., 2011). Since 2003, the government of China has been boosting the AD industry by direct financial ranging from 1000 million CNY (China Yuan) in 2003–2005 to 2500 million CNY in 2006–2007. The Renewal Energy Law was introduced in 2006 to drive the biogas engineering projects and the financial aid reached 5000 million CNY in 2010 (National People’s Congress NPC, 2005; Ministry of Agriculture MOA, 2007). This was instrumental in boosting an increase in AD projects from 2300 in 2003 to 10,000 in 2013 (Gu et al., 2016). After 2009, China increased its to the AD industry through subsidies covering up to 45% of the total project cost, along with introducing FITs. To improve the performance of existing plants, local biogas service systems were set up. China has a target of achieving 44 Giga cubic meters biogas production by 2020 of which a minimum household share of 30 Giga cubic meters is expected (National Development and Reform Commission NDRC, 2007).

16.3.6.7 India

A rural energy crisis along with the rising cost of energy imports has drivev the attention of Indian policymakers toward the freely available energy resources such as agriculture and animal waste. In 1981, the introduction of the first biogas development program was instrumental in leading to the installation of five million family biogas plants (CSO and Office, 2014) in India. In addition, 400 biogas off-grid power plants have been established, with a power-yielding capacity of about 5.5 MW (MNRE, 2015). The current Indian policies in favor of the AD industry include the target of reduction of GHG emissions, and regulations restricting soil and water pollution—those related to human health (Mittal et al., 2018). An innovative approach to cleanliness was introduced by

the government under the name of Clean India Mission, under which financial assistance of up to 35% of the project cost is provided to waste treatment technologies (Breitenmoser et al., 2019). In line with these campaigns, composting and AD technology are recognized as the only viable municipal waste treatment methods (Breitenmoser et al., 2019). The Ministry of New and Renewable Energy (MNRE) has set a target of establishing 100,000 biogas plants by 2022 under the National Biogas and Manure Management Program (NBMMP) (MNRE, 2017). This scheme provides subsidies for the installation of family-size biogas systems. The key objective is to utilize organic waste from agriculture, animal husbandry, and household waste in rural and semiurban areas for biogas generation, which in turn would be used as a clean cooking fuel and source of lighting.

16.3.6.8 Others

Key policies applied in Denmark, Austria, and the Netherlands to biogas production are listed in Table 16.2.

Table 16.2

Country 1993 FIT policy

Policy Constant rate of 4.3 €ct/kWh for AD industry

1999 Renewable portfolio standard with a system o 2004 FIT policy and its amendment in 2009

The Green Growth Agreement, 2009 Energy Strategy, 2050 (2011) Estrada et al. (2012) 1998 FIT policy (amended 2002, 2006, 2010)

Constant rate of 5 €ct/kWh for 13 years (Bangalore

National Green Electricity act, 2002 (amended 2006 2003 MEP (Environmental Quality of Electricity Production)

7.7 €ct/kWh for 10 years, biogas production increas 2006 SDE (new stimulation program) 2011 SDE+

16.4 Decision- system for biomethane implantation with techno-economic analysis and policies

There is a plethora of reports available on the techno-economic evaluation of various fuels; however, an inclusive database incorporating techno-economic evaluation at the supply chain level of fossil and renewable energy sources is lacking. To that end, an Excel-based mathematical model, called the Decision for Digester-Algae IntegRation for Improved Environmental and Economic Sustainability (DAIRIEES), was developed incorporating technical and economic data (INL-USDENL, 2017). DAIRIEES quantified mass and nutrient transport among the different subsystems of a plant and provided information on carbon and nutrient sequestration, output and market value of products, and overall economic feasibility. There are various challenges faced when developing a decision- system. One of these is the variety of technologies used in AD. Another issue is the existence of different biogas generation systems which entail comparable expenses. Lastly, one of the major challenges is the need to use economic, technical, and environmental information to make an integrated analysis. A multicriteria tool for technoeconomic analysis of biomethane generation was developed by Billig and Thrän (2016), while Estrada et al. (2012) focused on the sensitivity analysis of design parameters and product cost of odor mitigation technology. Both studies reported a lower sensitivity and costs when using biological methods in comparison to physico-chemical methods. However, there is a requirement for comprehensive tools which are able to organize the decision-making process and aid the deciding committee to choose the most suitable option among different biomethane generation technologies. To address the bottlenecks associated with individual approaches mentioned above, Kuznetsova et al. (2019) proposed an integrated decision- methodology (DSM) incorporating waste modeling, WtEMS optimization, and multidimensional assessment frameworks. The proposed DSM by the author was tested on a case study that was located in Singapore. A total of a 50% reduction in operational costs decreased land use by 74.8%, reduced transportation by 15.3%, and also increased the revenue from

electricity two-fold. It is imperative that a decision- system is developed based on multicriterion information, including technical, social, and environmental.

16.5 Conclusion

To commercialize biomethane technologies, policy is necessary. Some of the successful policies implemented in Europe could be used in other countries as well for efficient commercialization. Dynamic through governmental policies can help biomethane technologies to reach market maturity. Most of the renewables compete with established fossil energy sources, where the role of government through policies is pivotal in fighting climate change. A strong policy will encourage the industry toward innovation, thus reducing the cost. An integrative decision- system is necessary to assess the practical outreach of these policies and technology implementation. This chapter covered the techno-economic and policy implications of biomethane systems in various countries. However, there are certain other factors which need to be addressed in the future, including the social dimensions and a market analysis. For example, heat after power generation or waste heat could be used for district heating, which could bring in additional revenues. This possibility is limited in tropical countries as it requires cooling due to their climate. Similarly, social acceptance is necessary, such as sorting the waste, and using the biomethane from waste for cooking. Hence, a summative model is necessary to evaluate these multidimensional factors.

References

Adelt et al., 2011 Adelt M, Wolf D, Vogel A. LCA of biomethane. J Nat Gas Sci Eng. 2011;3:646–650. Akay et al., 2005 Akay G, Dogru M, Calkan OF, Calkan B. Biomass processing in biofuel applications. In: Lens P, ed. Biofuels for Fuel Cells: Renewable Energy From Biomass Fermentation. London: IWA; 2005;51–75. Baccioli et al., 2019 Baccioli A, Ferrari L, Guiller R, Yousfi O, Vizza F, Desideri U. Feasibility analysis of bio-methane production in a biogas plant: a case study. Energies. 2019;12(3):473 https://doi.org/10.3390/en12030473/. Bangalore et al., 2016 Bangalore M, Hochman G, Zilberman D. Policy incentives and adoption of agricultural anaerobic digestion: a survey of Europe and the United States. Renew Energy. 2016;97:559–571. Bartoli et al., 2019 Bartoli A, Hamelin L, Rozakis S, Borzęcka M, Brandão M. Coupling economic and GHG emission ing models to evaluate the sustainability of biogas policies. Renew Sustain Energy Rev. 2019;106:133–148. Billig and Thrän, 2016 Billig E, Thrän D. Evaluation of biomethane technologies in Europe—technical concepts under the scope of a Delphi-survey embedded in a multi-criteria analysis. Energy. 2016;114(C):1176–1186. Bracmort, 2010 Bracmort, K. Anaerobic Digestion: Greenhouse Gas Emission Reduction and Energy Generation, 2010, May 4. Braun et al., 2008 Braun R, Weiland P, Wellinger A, et al. Biogas from energy crop digestion. IEA Bioenergy Task 2008;37. . Breitenmoser et al., 2019 Breitenmoser L, Gross T, Huesch R, et al. Anaerobic digestion of biowastes in India: opportunities, challenges and research needs. J

Environ Manag. 2019;236:396–412. Bundgaard et al., 2014 Bundgaard, S.S., Kofoed-Wiuff, A., Herrmann, I.T. and Karlsson, K.B., 2014. Experiences with biogas in Denmark. Bušić et al., 2018 Bušić A, Kundas S, Morzak G, et al. Recent trends in biodiesel and biogas production. Food Technol Biotechnol. 2018;56(2):152–173. Capodaglio et al., 2016 Capodaglio AG, Callegari A, Lopez MV. European framework for the diffusion of biogas uses: emerging technologies, acceptance, incentive strategies, and institutional-regulatory . Sustainability. 2016;8(4):298. CSO and Office, 2014 CSO and Office. In: Office CS, ed. Energy Statistics. New Delhi: Ministry of Statistics and Programme Implementation; 2014. Cucchiella et al., 2018 Cucchiella F, D'Adamo I, Gastaldi M, Miliacca M. A profitability analysis of small-scale plants for biomethane injection into the gas grid. J Clean Prod. 2018;184:179–187. Damave, 2018 Damave M.-C. (2018) Key points of the agr’iDay of 03/07/2018: innovating for the future: bioeconomy in the UK and . 11 p. DENA, 2011 DENA. The role of natural gas and biomethane in the fuel mix of the future in , 2011. Drosg, 2011 Drosg, B. 2011. Austrian Country Report, 2011. Presented at IEA Bioenergy Task 37 Meeting. Cork, Ireland 14–16 September 2011. EC, 2018 EC, 2018/844. DIRECTIVE (EU). . Edwards et al., 2015 Edwards J, Othman M, Burn S. A review of policy drivers and barriers for the use of anaerobic digestion in Europe, the United States and Australia. Renew Sustain Energy Rev. 2015;52:815–828. EREC, 2009 EREC, Renewable Energy Policy Review e Denmark, EREC e European Renewable Energy Council, Brussels, Belgium, 2009. Estrada et al., 2012 Estrada JM, Kraakman NJRB, Lebrero R, Muñoz R. A

sensitivity analysis of process design parameters, commodity prices and robustness on the economics of odour abatement technologies. Biotechnol Adv. 2012;30(6):1354–1363. European Parliament, 2009 European Parliament, 2009. Directive 2009/28/EC of the European Parliament and of the Council. Off. J. Eur. Union 62. Eurostat, 2012 Eurostat, Supply, transformation, Consumption e Renewables and Wastes (Total, Solar Heat, Biomass, Georthermal, Wastes) e Annual Data, 2012. Accessed June 2012, . Fachverband Biogas, 2017 Fachverband Biogas, 2017. German Biogas Association. Fan et al., 2018 Fan YV, Klemes JJ, Lee CT, Perry S. Anaerobic digestion of municipal solid waste: energy and carbon emission footprint. J Environ Manag. 2018; https://doi.org/10.1016/j.jenvman.2018.07.005. Ferella et al., 2017 Ferella F, Puca A, Taglieri G, Rossi L, Gallucci K. Separation of carbon dioxide for biogas upgrading to biomethane. J Clean Prod. 2017;164:1205–1218. Ferreira et al., 2012 Ferreira M, Marques IP, Malico I. Biogas in portugal: status and public policies in a european context. Energy Policy. 2012;43:267–274 http://doi.org/10.1016/j.enpol.2012.01.003. Fubara et al., 2018 Fubara T, Cecelja F, Yang A. Techno-economic assessment of natural gas displacement potential of biomethane: a case study on domestic energy supply in the UK. Chem Eng Res Des. 2018;131:193–213. Gabauer et al., 2008 Gabauer, W., Waltenberger, R., Neureiter, M., Kirchmayr, R., Braun, R. 2008. Biogas in Austria Facts and Figures, Energy Crops, Technologies (National Seminar, Jyvaskya, Finland). Georgakakis et al., 2003 Georgakakis D, Christopoulou N, Chatziathanassiou A, Venetis T. Development and use of an economic evaluation model to assess establishment of local centralized rural biogas plants in Greece. Appl Biochem Biotechnol. 2003;109:275–284.

Gu et al., 2016 Gu L, Zhang YX, Wang JZ, Chen G, Battye H. Where is the future of China’s biogas? Review, forecast, and policy implications. Pet Sci. 2016;13(3):604–624. Hermann and Hermann, 2018 Hermann, L., Hermann, R. 2018 Report on regulations governing AD and NRR in EU member states. Systemis: circular solutions for biowaste. The Netherlands, 123 p. IEA, 2009 IEA, 2009. Renewable energy sources act 2009. . IEA, 2011 IEA, 2011. Feed-in Programme SDE + (Stimulering Duurzame Energie +) act 2011 . IEA, 2012 IEA, 2012. Renewable energy sources act 2012. . INL-USDENL, 2017 INL-USDENL, Decision- for Digester-Algae IntegRation for Improved Environmental and Economic Sustainability (DAIRIEES) Manual, INL U.S. Department of Energy National Laboratory, INL/EXT-14-32566, 2017/. Jiang et al., 2011 Jiang X, Sommer SG, Christensen KV. A review of the biogas industry in China. Energy Policy. 2011;39(10):6073–6081. Khan et al., 2017 Khan IU, Othman MHD, Hashim H, et al. Biogas as a renewable energy fuel—a review of biogas upgrading, utilisation and storage. Energy Convers Manag. 2017;150:277–294 https://doi.org/10.1016/j.enconman.2017.08.035. Kuznetsova et al., 2019 Kuznetsova E, Cardin A-M, Diao M, Zhang S. Integrated decision methodology for combined centralized-decentralized waste-to-energy management systems design. Renew Sustain Energy Rev. 2019;103:477–500 ff10.1016/j.rser.2018.12.020ff. ffhal-02184298f. Larsson et al., 2016 Larsson M, Grönkvist S, Alvfors P. Upgraded biogas for transport in Sweden-effects of policy instruments on production, infrastructure deployment and vehicle sales. J Clean Product. 2016;112

http://doi.org/10.1016/j.jclepro.2015.08.056. Lee, 2017 Lee DH. Evaluation the financial feasibility of biogas upgrading to biomethane, heat, CHP and AwR. Int J Hydrog Energy. 2017;42:27718–27731. Le Fevre, 2014 Le Fevre, C., 2014. The prospects of natural gas as a transport fuel in Europe. Miltner et al., 2017 Miltner M, Makaruk A, Harasek M. Review on available biogas upgrading technologies and innovations towards advanced solutions. J Clean Prod. 2017;161:1329–1337. Ministry of Agriculture (MOA), 2007 Ministry of Agriculture (MOA). Development plan for the agricultural bioenergy industry 2007–2015. 2007. (in Chinese). MISE, 2018 MISE, 2018. Interministerial Decree of 2 March 2018. Promotion of the use of biomethane 706 and other advanced biofuels in the transportation sector. Available online: 707 http://www.sviluppoeconomico.gov.it/index.php/en/. Mittal et al., 2018 Mittal S, Ahlgren EO, Shukla PR. Barriers to biogas dissemination in India: A review. Energy Policy. 2018;112:361–370. MNRE, 2015 MNRE, 2015. Annual Report, 2015–16. In: Energy, M.O.N.A.R. (Ed.), New Delhi. MNRE, 2017 MNRE, 2017. Ministry of New and Renewable Energy. Annual Report 2016-2017. Government of India. . MNRE, 2019 MNRE, 2019. . Murphy and Power, 2008 Murphy JD, Power N. Technical and economic analysis of biogas production in Ireland utilizing three different crop rotations. Appl Energy. 2008;10:3–15. Nachmany et al., 2015 Nachmany M, Fankha S, Davidová J, et al. The 2015 global climate legislation study. A Rev Clim change legislation in 2015;99.

National Development and Reform Commission (NDRC), 2007 National Development and Reform Commission (NDRC). Medium and long-term development plan for renewable energy in China. 2007. (in Chinese). National People’s Congress (NPC), 2005 National People’s Congress (NPC). Renewable energy law of the People’s Republic of China. 2005. (in Chinese). Niesner et al., 2013 Niesner J, Jecha D, Stehlik P. Biogas upgrading techniques: state of art review in european region. Chem Eng Trans. 2013;35:517–522. OGEM, 2017 OGEM, 2017. Tariffs and Payments: Non-domestic RHI. . Pellegrino et al., 2017 Pellegrino S, Lanzini A, Leone P. Greening the gas network—the need for modelling the distributed injection of alternative fuels. Renew Sustain Energy Rev. 2017;70:266–286 https://doi.org/10.1016/j.rser.2016.11.243. Poliafico and Murphy, 2007 Poliafico, M., Murphy, J.D., 2007. Anaerobic digestion in Ireland: decision system, Department of Civil, Structural and Environmental Engineering. Cork Institute of Technology, Ireland, 2007. Ravina and Genon, 2015 Ravina M, Genon G. Global and local emissions of a biogas plant considering the production of biomethane as an alternative end-use solution. J Cleaner Prod. 2015;102:115–126. Ropenus and Jacobsen, 2015 Ropenus, S., Jacobsen, H.K., 2015. A Snapshot of the Danish Energy Transition: Objectives, Markets, Grid, Schemes and Acceptance. Study. Rotunno et al., 2017 Rotunno P, Lanzini A, Leone P. Energy and economic analysis of a water scrubbing based biogas upgrading process for biomethane injection into the gas grid or use as transportation fuel. Renew Energy. 2017;102:417–432. Scarlat et al., 2015 Scarlat N, Dallemand JF, Monforti-Ferrario F, Nita V. The role of biomass and bioenergy in a future bioeconomy Policies and facts. Env

Dev. 2015;15:3–34 http://doi.org/10.1016/j.envdev.2015.03.006. Shea et al., 2017 Shea R, Wall DM, Kilgallon I, Browne JD, Murphy JD. Assessing the total theoretical, and financially viable, resource of biomethane for injection to a natural gas network in a region. Appl Energy. 2017;188:237–256. Swedish Environmental Protection Agency, 2017 Swedish Environmental Protection Agency, 2017. Swedish environmental goals concerning food waste sorting. . UK DECC, 2015 UK DECC. Renewable Heat Incentive (RHI) United Kingdom: Department of Energy & Climate Change; 2015. USDA, 2014 USDA (2014) Biogas opportunities roap. U.S. Department of Agriculture, U.S. Environmental Protection Agency, U.S. Department of Energy, 27 p. Uusitalo et al., 2015 Uusitalo V, Havukainen J, Soukka R, Väisänen S, Havukainen M, Luoranen M. Systematic approach for recognizing limiting factors for growth of biomethane use in transportation sector—a case study in Finland. Renew Energy. 2015;80:479–488. Vasco-Correa et al., 2018 Vasco-Correa J, Khanal S, Manandhar A, Shah A. Anaerobic digestion for bioenergy production: g lobal status, environmental and techno-economic implications, and government policies. Bioresour Technol. 2018;247:1015–1026. Vo et al., 2018 Vo TTQ, Wall DM, Ring D, Rajendran K, Murphy JD. Technoeconomic analysis of biogas upgrading via amine scrubber, carbon capture and ex-situ methanation. Appl Energy. 2018;212:1191–1202.

Chapter 17

Large-scale biogas upgrading plants: future prospective and technical challenges

Ram Chandra Poudel¹, Dilip Khatiwada², Prakash Aryal³ and Manju Sapkota⁴, ¹1Department of Biological Sciences, University of Bergen, Bergen, Norway, ²2Division of Energy Systems, Department of Energy Technology, KTH Royal Institute of Technology, Stockholm, Sweden, ³3Department of Chemical Engineering, Monash University, Clayton, VIC, Australia, ⁴4Institute of Chemistry, Bioscience and Environmental Engineering, Faculty of Science and Technology, University of Stavanger, Stavanger, Norway

Abstract

Biogas is a clean and renewable energy sources. It can be used in different enduse applications such as transport fuel, electricity generation, heating, and cooking. It is produced from the anaerobic digestion of organic matter/waste. Biogas needs to be upgraded into biomethane for its use as a transport fuel and injection into the natural gas grid. In this chapter, the state-of-the-art of the biogas upgrading technologies, especially biological ones implemented in largescale biogas plants, are analyzed and discussed. This includes different process parameters and conditions for the downstream upgrading of biogas into biomethane. Biogas upgrading into biomethane and its utilization could help in decarbonizing the transport sector. Similarly, the injection of biomethane into the gas grid might enhance energy security, reduce fossil energy consumption, and contribute to economic development.

Keywords

Biogas upgrading technologies; biomethane; transport fuel; gas grid

Chapter outline

Outline


17.2 Biogas composition and feedstock types 468

17.3 Biogas upgrading for natural gas grid injection and transport fuel 468

17.4 State-of-the-art of large-scale biogas upgrading technologies 472

17.4.1 Physicochemical upgrading technologies 472 17.4.2 Power-to-gas technology for methanation 474 17.4.3 Bioelectrochemical system (Cambrian Innovation) 482 17.4.4 Photosynthetic biogas upgrading system 484

17.5 Conclusion and future perspective 485

References 486

17.1 Introduction

Biogas is mainly composed of methane (CH4) and carbon dioxide (CO2), and is produced from the degradation of organic matters in the absence of oxygen. It also contains small amounts of other gases such as hydrogen sulfide (H2S), oxygen (O2), ammonia (NH3), and siloxanes. Generally, the calorific value of raw biogas lies in the range of 20–30 MJ/m³ depending on the composition of the gas produced (Perez-Sanz et al., 2019). Predominately, raw biogas is directly combusted for electricity generation at combined heat and power plants (CHPs) and for cooking purposes. However, other gaseous components including CO2 should be removed when it is used on a large scale such as for vehicle fuel and/or injected into the natural gas grid (Luo and Angelidaki, 2012; Ryckebosch et al., 2011). Therefore the upgrading of biogas is required for its utilization as a transport fuel and grid injection. Biomethane could be one of the potential vehicle fuels for meeting renewable energy targets, for example, 14% of energy consumption in the transport sector in the EU, and it could be injected into the natural gas grid. Several biogas cleaning or upgrading technologies have been commercialized such as water scrubbing, pressure swing adsorption (PSA), membrane separation, chemical absorption, cryogenic, and biological technologies (Rodero et al., 2018). However, bioelectrochemical system (BES), power-to-gas (PtG), hydrogen mediated microbial biogas upgrading, and hybrid technologies are in the development stages, mostly in laboratory-scale reactors. Such laboratoryscale reactors and upgrading technologies should be optimized for CO2 utilization and removal performances while retaining low cost for CH4 enrichment (Angelidaki et al., 2018). This chapter extensively describes the scaling-up emerging biogas upgrading technologies, in particular hydrogenmediated and PtG technologies.

17.2 Biogas composition and feedstock types

Biogas is composed of CH4 (40%–75%) and CO2 (15%–60%) as the main constituents, while other gases like water vapor (H2O), nitrogen (N2), oxygen (O2), hydrogen sulfide (H2S), ammonia (NH3), hydrocarbons, siloxane, etc. are known as biogas impurities, and are present only in small amounts (Ryckebosch et al., 2011; Scholz et al., 2013). Different types of biomass from various sources can be utilized as a feedstock for the production of raw biogas. This can be broadly categorized into five different groups, namely, agricultural wastes, municipal solid wastes, wastewater, industrial wastes, and landfill wastes, as shown in Fig. 17.1 (IEA Bioenergy, 2017). Primarily, the composition of the produced biogas heavily relies on the feedstock composition and its characterization, whereas the pretreatment options, digestion system design, trace metal concentrations, microbial composition, etc. contribute to slight variations (Demirel and Scherer, 2011; Joo et al., 2018; Mshandete et al., 2006). Depending on the proportion of constituents, the quality of the biogas varies and its utilization may have adverse effects on appliances (e.g., corrosion), human health (e.g., toxicity), and the environment (pollution), as presented in Table 17.1. Therefore it is crucial to remove all the impurities along with the CO2 from the biogas mixture. It helps to protect from possible undesirable effects and to increase the concentration of CH4, which eventually enhances the calorific value and relative density of the biogas.

Figure 17.1 Different types of feedstock for large-scale biogas plants (IEA Bioenergy, 2017).

Table 17.1

Biogas constituents

Biogas (biomass)

Biogas (waste water/sludge)

Biogas (landfill waste)

CH4 (vol.%)

60–70

55–75

35–65

CO2 (vol.%)

30–40

19–45

15–40

N2 (vol.%)

0.2

0–2

15

O2 (vol.%)

0

0–1

1

H2 (vol.%)

0

Nf

0–3

Hydrocarbons total (vol.%)

0

Nf

0

NH3 (ppm)

100

Nf

5

H2S (ppm)

0–4000

63–3000

0–100

Siloxanes (ppmv)

0–0.4

1.5–15

0.1–4

H2O vapor (mg/N m³)

1–5a

a

1–5a

Total chlorine (mg/N m³)

100

Nf

5

AD, Anaerobic digestion; CHP, combined heat and power plant; m³, cubic meters; mg, milligrams; Nf, not found; ppm, parts per million by weight; ppmv, parts per million by volume; Wobbe index, energy output (during combustion) of biogas or biomethane.

aSaturation at biodigester temperature.

17.3 Biogas upgrading for natural gas grid injection and transport fuel

Biogas is used as an alternative renewable fuel in CHP plants for electricity and heat production and in domestic stoves for residential cooking (Aryal and Kvist, 2018; Aryal et al., 2018). The applications are further extended to substitute natural gas either by injecting into the gas grid system that provides an unlimited storage and distribution system for domestic uses, cogeneration plant, and industry, or by supplying it as a transport fuel (Bailón and Hinge, 2012). However, biogas needs to be upgraded into biomethane to meet the standards and purity requirement of both gas grid injection and as a transport fuel. The biomethane parameters and the standards set in different countries compared with natural gas quality are shown in Table 17.2.

Table 17.2

Parameters Lower heating value (MJ/N m³)

Sweden (vehicle fuel) (Awe et al., 2017; Bailón and Hinge, 2012) Nf

WI Higher (H)

44.7–46.4

Lower (L)

43.9–47.3

Methane (vol.%)

≥97

CO2 (vol.%)

≤3

O2 (vol.%)

≤1

N2 (vol.%)

Nf

Heavy hydrocarbons (vol.%)

Nf

H2 (vol.%)

≤0.5

H2S (ppmv)

≤15.2

Total sulfur (ppmv)

<23

NH3 (ppmv)

≤20

Mercaptans (ppmv)

Nf

Halogenated compounds (mg/N m³)

≤1

Siloxanes (ppm)

Nf

H2O dew point (°C)

≤−5 at AT ≤−9 at 200 bar

Nf, Not found; WI, Wobbe index; AT, atmopheric temperature.

17.4 State-of-the-art of large-scale biogas upgrading technologies

Biogas upgrading technologies are often multistaged process involving removal of trace components such as H2S, H2O, and siloxanes, and separation and/or conversion of CO2 to produce biomethane at the set standards. Different upgrading technologies are summarized in the following sections.

17.4.1 Physicochemical upgrading technologies

Physicochemical upgrading technologies are conventional technologies which utilize different the chemical and physical properties of the gases to be removed and may undergo further chemical reaction (Rodero et al., 2018). Depending on the chemophysical mechanisms involved in the impurities to be removed, these technologies can be grouped further into three sub-categories, namely (1) absorption (water, physical, and chemical), (2) adsorption (pressure swing), and (3) separation (membrane and cryogenic technology) (Adnan et al., 2019). Commonly used absorption techniques are water scrubbing, organic physical absorption (OPA), and chemical (amine) absorption. They are differentiated with each other only by the types of absorbents used in the column. However, the absorbent increases the mass-liquid transfer in each absorption system (Adnan et al., 2019). Water scrubbing is the most widely adopted physicochemical upgrading technology, as shown in Fig. 17.2. In water scrubbing, CO2, which is more highly soluble in water than CH4 at low temperature and high pressure, is absorbed, leaving the CH4. Therefore water leaving the column is saturated with CO2 and thus gets separated (Scholz et al., 2013). The OPA technique is based on an absorption mechanism similar to water scrubbing but organic solvents such as selexol and genosorb are used instead of water (Angelidaki et al., 2018; Muñoz et al., 2015). In the chemical absorption process, a chemical reaction

takes place between the solvent [e.g., monoethanolamine, diethanolamine (DEA), or diglycolamine (DGA)] and the absorbed substance, that is, CO2 present in the biogas where the concentrated CH4 is left behind and separated from the CO2 (Awe et al., 2017; Niesner et al., 2013). However, the latter two (OPA and chemical) absorption processes demand extra cost to regenerate the used solvent compared to the former one (Niesner et al., 2013).

Figure 17.2 Number of physicochemical upgrading technologies in 2017 ( EBA, 2018).

Similarly, the PSA technique is a dry method that separates CO2 and CH4 based on the physical properties of the gases and adsorbents (Bauer et al., 2013). In this process, compressed raw biogas is injected into an adsorption column containing adsorbents, namely, silica gel, zeolites, and activated carbon (Adnan et al., 2019; Bauer et al., 2013). Due to the smaller size of CO2 than CH4, it gets adsorbed on the adsorbents by van der Waals forces, thereby retaining CO2 in adsorbents (Adnan et al., 2019; Prussi et al., 2019). Membrane separation technology is one of the common types of upgrading process. It is based on the separation of gases by a selective membrane, where CO2 es through it while CH4 is retained by the membrane. This process highly depends upon the type of membrane (e.g., cellulose acetate, silicone rubber, hollow fibers, etc.) and the pressure applied (Scholz et al., 2013; Scholz et al., 2015). The pressure of the upgraded biogas is equivalent to the natural gas grid injection pressure and therefore further treatment is not necessary as the gas grid injection is the end-use (Scholz et al., 2013). Cryogenic separation is an emerging technology in the biogas upgrading process in which the gas products are separated based on their condensation temperature difference, that is, the boiling points of CO2 and CH4 are −78°C and 160°C, respectively (Awe et al., 2017; Thrän UFZ et al., 2014). When the gas mixture is cooled at −85°C to −110°C, CO2 is sublimated and separated in solid form (Miltner et al., 2017). Other gases such as O2, siloxanes, and N2, can also be separated by the temperature condensation and distillation process (Awe et al., 2017). The separated liquid CO2 can be utilized as a raw material for a separate process. Therefore this technique is more environmentally friendly (Yousef et al., 2016). However, the methane loss during the upgrading process can be reduced by the application of a regenerative thermal oxidizer (Kvist and Aryal, 2019). Table 17.3 provides the technical parameters of various physicochemical biogas upgrading processes.

Table 17.3

Technology

CH4 recovery (%)

CH4 loss (%)

H2S (ppmv)

Water scrubbing

>97

<2

<2

Organic physical (solvent) scrubbing

96–98.5

<2

7–8

Chemical (amine) scrubbing

>99

<0.05

<4

Pressure swing adsorption

98

2–4

–

Membrane separation

99.5 (MSS) 90–97 (SSS)

<0.5

–

Cryogenic separation

90–98

<2

–

MSS, Multistage system; SSS, single-stage system.

17.4.2 Power-to-gas technology for methanation

PtG is a novel technology and is used for biomethane production by converting CO2 into methane. Some European countries, including and Denmark, have periodical surplus electricity from wind turbines due to seasonal variations in the weather. Therefore this surplus electricity can be utilized to generate H2 using a water electrolysis process. H2 thus produced through renewable resources reacts further with the CO2 fraction of the biogas to produce biomethane. Therefore PtG technology utilizes surplus and seasonal low-cost electricity for biomethane production, thereby contributing to power grid balancing and decarbonizing fossil-based natural gas grids (Heide et al., 2010; Kötter et al., 2015). Moreover, the PtG technology combined with biogas upgrading technology can reduce the biomethane-specific carbon footprint and increase the methane concentration utilizing CO2 (Miltner et al., 2016). Injected biomethane-rich gas can be reconverted into electricity to meet household or commercial electricity demand and/or vehicle fuels (Collet et al., 2017; Jentsch et al., 2014). The application of PtG methanation upgrading technology is practicable where the surplus electricity generated from renewable sources can be used to produce H2 (Hoyer et al., 2016). Moreover, the efficiency and sustainability of the overall process of PtG combined with biogas upgrading largely depends upon the availability of electricity for electrolysis, system design, operating voltage, reaction mechanism, temperature, pressure, catalyst availability, catalyst promoter, catalyst degeneration and regeneration, microbial growth rates, and their operating parameters, among other factors (Blanco and Faaij, 2018; Frontera et al., 2017; Gahleitner, 2013). Nevertheless, the overall process cost is largely determined by the electrolysis process since it is a more expensive process than methanation (Gotz et al., 2016; Thema et al., 2019a). Through the process of

proper heat integration, the heat generated during the hydrogenation/methanation process can be part of an electrolysis process which improves the exergy efficiency of the integrated plant (Gotz et al., 2016; Mendoza-Hernandez et al., 2019). The most common electrolyzers are proton exchange membrane, also known as polymer electrolyte membrane (PEM), and alkaline electrolyzer cell, while other types include solid oxide electrolyzer cell (SOEC) and alkaline solid polymeric membrane electrolyzer (alkaline-SPE) (Eveloy and Gebreegziabher, 2018; Wulf et al., 2018). The PtG for methanation can be either done by a chemical catalytic or biological process (Sections 17.4.2.1 and 17.4.2.2).

17.4.2.1 Catalytic/thermochemical methanation

The catalytic methanation process takes place at high pressure (<100 bar) and elevated temperature (200°C–550°C) in the presence of a metal catalyst (Schaaf et al., 2014; Schwede et al., 2017; Xia et al., 2016). The thermodynamic reactions and selection of the catalysts such as nickel, copper, rhodium, ruthenium, and palladium have a significant effect on the activation and reduction steps of the CO2 methanation process (Frontera et al., 2017; Gao et al., 2015). The reactor configuration is one of the critical parameters due to the exothermic nature of the reaction taking place during the process where twophase fixed-bed reactors (FBRs) and fluidized bed reactors are predominately used (Schaaf et al., 2014). The methanation process significantly contributes to heat regeneration and greater space–time yields (Eveloy and Gebreegziabher, 2018; Thema et al., 2019a, 2019b). Different catalytic upgrading technologies including pilot, demonstration, and commercial scales, with the technical specifications, are described in Table 17.4.

Table 17.4

Catalytic methanation

Scale

Type

EI (kWe)

Reactor type

Electrolyzer

CH4 content (%)

ETOGAS and Audi e-gas

Commercial

Isothermal fixed bed

Alkaline

6000

MeGa-StoRE

Pilot

Air cooled

–

Bottled H2

Haldor Topsøe

Pilot

Reactor with steam

SOEC

50

RENOVA GAS

Demonstration

Multichannel

Alkaline

15

HELMETH

Pilot

Isothermal

SOEC

30

EI, Electricity input; MeGa-StoRE, methane gas storage of renewable energy; SOEC, solid oxide electrolyzer cell; Tempr, temperature.

A few selected plants of catalytic methanation, which are in the process of commercialization, are described below.

ETOGAS and Audi e-gas technology

This is one of the catalytic methanation technologies that have been implemented in three biogas plants in . The first two are pilot plants with capacities of 25 and 250 kW (input electricity) located in Bad Hersfeldt and Stuttgart, respectively. The third one is a t project between ETOGAS and Audi of capacity 6 MW input electricity (three, each with 2 MW units) in Wertle. The source of carbon for this process is nonupgraded biogas for pilot scale, air for the demonstration plant in Stuttgart, and separated CO2 from the upgrading plant. ETOGAS claimed to have synthetic methane (Audi e-gas) above 96% (Bailera et al., 2017; Benjaminsson et al., 2013). In the electrolysis unit, the electrolyzer used is an alkaline electrolyzer with nickel-based catalyst as well as electrodes. Similarly, the methanation reactor is made up of a simple isothermal FBR with a capacity of 325 N m³/h. When required, the capacity can be increased by installing additional reactors (Benjaminsson et al., 2013).

Haldor Topsøe

In 2013, a proof-of-concept based on catalytic methanation technology was

installed by Haldor Topsøe in Foulum, Denmark. This pilot plant used 40 kW SOEC. The PtG exergy efficiency of the plant was 79.8% and the CO2 in the biogas was upgraded to gas grid quality requirements (Benjaminsson et al., 2013). Similarly, Haldor Topøse installed another three adiabatic FBR pilot plants at Aarhus University, Denmark, which were operated for 1000 h. The first two reactors were designed to remove sulfur contents (<0.04 ppmv), while the last reactor was for the removal of other impurities. For methanation to occur, H2 was injected into the third reactor (Bailera et al., 2017; Benjaminsson et al., 2013).

Methane gas storage of renewable energy

Methane gas storage of renewable energy (MeGa-StoRE) is a flexible catalytic methanation project based on a single-step Sabatier process that is coupled with HyProvide electrolyzer modules, a gas purifying unit (H2S removal), and a methanation reactor (Bailera et al., 2017; ForskEL, 2016). The hydrogen produced from the electrolyzer (advanced alkaline or PEM electrolyzer) is injected into the methanation reactor, which then reacts with CO2 derived from the gas purifying unit to form CH4 (ForskEl, 2015). As a proof-of-concept, MeGa-StoRE1 methanation was operated in 2015, at higher temperature (270°C) and pressure (8 bar), which illustrated higher CO2 conversion efficiency up to 97%–99% and the CH4 production rate was 1 N m³/h (with 35% CO2 input) (ForskEL, 2015). The benefits of this technology include higher CO2 conversion efficiency, better yield of the biomethane, and compact design that led to a low carbon footprint (ForskEL, 2016). The MeGa-StoRE projected a milestone 0.25 MW methanation demonstration plant with capacity 43 N m³/h CH4 (with 35% CO2 input) by 2020 and 10 MW large upgrading system with capacity 1710 N m³/h CH4 (with 35% CO2 input) by 2035 and 2050 (ForskEL, 2015).

17.4.2.2 Chemoautotrophic (biological) methanation

The biological upgrading technique includes microbial, photosynthetic, and bioelectrochemical (electromethanogenesis) biogas upgrading and has been experimentally demonstrated at lab scale and demonstration-scale reactor (Angelidaki et al., 2018; Marín et al., 2018). This technology offers the unique opportunity to utilize CO2 from biogas as a feedstock either for biomethane production or biogas production. As a result, it can be considered as a carbonneutral technology. The microbial methanation is a cost-effective, low-energy, and environmentally friendly process. Nevertheless, microbial biogas upgrading is yet to be fully commercialized due to practical challenges associated with CO2 conversion efficiency and operating parameter optimization (Muñoz et al., 2015; Prussi et al., 2019). In a chemoautotrophic process, hydrogenotrophic methanogens act as biocatalysts, which utilize CO2 as a carbon source and electron acceptor, and H2 as an energy source and electron donor, as illustrated in Eq. 17.1 for CH4 production (Strevett et al., 1995; Voelklein et al., 2019). This process takes place under normal temperature and pressure conditions (Schwede et al., 2017; Strübing et al., 2017; Wahid et al., 2019). Likewise, the overall methanation process highly depends upon the nutrient supply, hydrogenotropic methanogen consortium, bioreactor configuration, availability of electron donor, that is, H2, and operational parameters (pH, temperature, etc.) (Bassani et al., 2017; Díaz et al., 2015; Treu et al., 2018; Zabranska and Pokorna, 2018). So far, different bioreactor types such as a continuous stirred tank reactor (CSTR), FBR/anaerobic filter, and trickle bed reactor have been tested for biomethanation (Lecker et al., 2017). The reactor design aims to mitigate the gas–liquid mass transfer limitations and to increase H2 solubility in the liquid phase during this process (Ullrich et al., 2018; Gotz et al., 2016). In this process, hydrogen is injected into the bioreactor for CO2 methanation. Therefore the cost of H2 production from a renewable source determines the cost of methanation (Vo et al., 2018). So far, the storage and transportation of H2 remain in the research and development (R&D) stage and are economically challenging (Niaz et al., 2015). In this context, PtG is practically relevant for on-site H2 production by utilizing surplus renewable energy (Rasmussen and Aryal, 2019).

(17.1)

CO2 biomethanation can be done by either injecting H2 directly into the anaerobic biodigester, called in situ, or using a separate bioreactor, known as the ex situ process (Zabranska and Pokorna, 2018), and a hybrid of these two designs is under further R&D (Angelidaki et al., 2018; Voelklein et al., 2019). In situ upgrading design is based on the injection of H2 inside the existing anaerobic digester where CH4 is produced by the reduction of CO2 by hydrogenotrophic methanogens using H2. The most common methanogens involved in this process are Methanococcus spp., Methanobacterium spp., Methanothermobacter spp., Methanoculleus spp., Methanosarcina spp., Methanospirillum spp., Methanococcus spp., and Methanomicrobium spp. (Bassani et al., 2015; Luo and Angelidaki, 2012; Rachbauer et al., 2017). The operational parameters like pH, temperature, mixing rate, and nutrient supply are crucial for the effectiveness of methanogenic bacteria. Therefore, when such parameters are fully controlled during the process, it will give rise to CH4 with optimal production (Luo et al., 2014; Strübing et al., 2017). As the process occurs within the existing biodigester, this design reduces the initial investment costs. During in situ upgrading, the addition of H2 inside the biodigester can increase the hydrogen partial pressure, causing volatile fatty acids accumulation, which finally may inhibit the digestion process by altering the pH beyond its optimum range (6.5–8.5) (Luo et al., 2012; Weiland, 2010). To overcome the problems faced by in situ design, such as shifting of microbial dynamics, ex situ upgrading is proposed. The initial steps of anaerobic digestion like hydrolysis and acidogenesis are absent in the ex situ upgrading process. Additionally, the bioreactor stability and efficiency of methane production depend upon CO2, H2, hydrogenotrophic archaea, and essential growth nutrients (Kougias and Angelidaki, 2018; Voelklein et al., 2019). As hydrogen is poorly soluble in water, it is not readily available to the methanogens. Therefore significant diffusion of H2 in the aqueous phase can be achieved by increasing the H2 pressure slightly, introducing an H2 diffusion appliance, and adjusting the gas circulation and mixing rate (Bassani et al., 2017), which ultimately boosts the H2 mass transfer rates and methanation process (Rodero et al., 2018; Seifert et al., 2014). In an experiment involving a separate thermophilic anaerobic reactor enriched with hydrogenotrophic methanogens, Luo and Angelidaki

(2012) found that the final output gas had 95% CH4 content at steady state when gas (mixture of biogas and hydrogen) was injected at a rate of 6 L/day. Similarly, the same percentage of methane content was observed by enhancing the stirring rate of the mixture in the reduced gas–liquid mass transfer condition. Likewise, Rachbauer et al. (2016) achieved complete CO2 conversion (>96%) with less than 0.1% residual H2 in the controlled ex situ bioreactor. In another research, Voelklein et al. (2019) noticed that when CO2 levels are below 9%, the CH4 generation rate from CO2 (in the presence of H2) is reduced significantly and thus recommended the hybrid technology combining in situ or ex situ units in series to solve this issue. Different types of chemoautotrophic methanation are under research and development, and MicrobEnergy and Electrochaea are the two examples of large-scale biogas upgrading technologies, as described in Table 17.5.

Table 17.5

Chemoautotrophic methanation

Scale

Reactor type

Electrolyzer

Technology

Energy input (kWe)

Temp (°C)

Pressure (bar)

Electrochaea (BioCat methanation)

Commercial

Ex situ

Alkaline

MicrobEnergy (BioPower2Gas)

Pilot

In situ+ex situ

PEM

PEM, Polymer electrolyte membrane.

Electrochaea

Electrochaea’s PtG technology is an ex situ biological methanation process, which involves H2 injection in a secondary bioreactor utilizing surplus electricity. Electrochaea’s BioCat methanation project is the first PtG technology to operate at a commercial-scale 1 MW plant by injecting biomethane into the Danish national gas grid (Electrochaea, 2019). This biological methanation uses an ex situ bioreactor where the efficient, robust, and pure strain of patented methanogens and their nutrient medium are inoculated. However, limited information is publically available. Before the commercialization of this technology, a precommercial project was designed and tested in a 10,000 L bioreactor using raw biogas as a CO2 source, which showed that Electrochaea’s PtG is highly efficient, stable, and productive for biomethanation. In addition, CO2 from biogas and H2 (produced by the electrolyzer) are supplied from the bottom where the methanation takes place at a temperature of 63°C and a pressure of 10 bar, resulting in a biomethane output above 99%. The biomethane can be directly injected into the gas grid and the recoverable heat produced can be utilized in on-site heating (Electrochaea, 2018).

MicrobEnergy

MicrobEnergy is another example of chemoautotrophic (biological methanation) which involves a two-reactor configuration: the first is to enhance the methane production by injecting H2 into an existing biogas reactor and the other is

designed for the methanation of pure CO2 using hydrogen. Therefore this technology involves an in situ and ex situ (hybrid) biogas upgrading system. The input power for electrolysis in the first biogas reactor is 120 kW, where the produced hydrogen is supplied from the bottom of the bioreactor to enhance gas–liquid mass transfer. This design significantly increased methane production by 25% compared to an untreated biogas mixture. Similarly, the second bioreactor is designed to reduce CO2 into biomethane by adding sufficient nutrients and hydrogenotrophic methanogens in a separate reactor. Hydrogen and CO2 are supplied from the bottom of the reactor to achieve better mixing of gas in the liquid phase. The temperature and pressure in the reactors are maintained at 40°C–65°C and 1 bar (atmospheric), respectively, and the methanation process remains stable by controlling the gas flow (Benjaminsson et al., 2013). Similarly, a demonstration plant (BioPower2Gas) research project was conducted in 2015 in Allendorf, , in 5 m³ CSTR. The operating conditions include temperature and pressure in the range of 50°C–80°C and 5– 15 bar, respectively. The methanation uses 30 m³ raw biogas as input at standard temperature and pressure and the methane content in the output gas was greater than 98% (IEA Bioenergy, 2018).

17.4.3 Bioelectrochemical system (Cambrian Innovation)

BES for biogas upgrading is an emerging technology where electro-active methanogens accept electrons from a charged cathode as an electron source, and CO2 from biogas acts as a carbon source for methane production. One of the leading companies, called Cambrian Innovation, commercialized BES technology for renewable electricity production from wastewater in microbial fuel cell-based technology. The company commercialized the first prototype EcoVolt Reactor to convert industrial wastewater into clean water and renewable methane gas with an efficiency of about 99.9% contaminants removal (Aryal et al., 2018). Nevertheless, limited information has been provided publically. The company further announced the BioVolt reactor prototype as a self-powered wastewater treatment system. Therefore such movement illustrated the upscaling

potential for BES technology to produce renewable methane and electricity (Aryal et al., 2018). The efficiency of the BES system relies on the electron transfer mechanism and transfer rate. The electrons are transferred from the externally poised cathode surface to microbes that occur either directly, indirectly, or via an electron mediator. Direct electromethanogenesis takes place via outer-membrane redox proteins, in particular c-type cytochromes, ferredoxin, rubredoxin, hydrogenase, and/or formate dehydrogenase, conductive pili, and nanowires. Mediated electromethanogenesis occurs through a different mediator such as bioelectrochemically produced H2, formate, acetate, flavins, riboflavins, quinones, and phenazines secreted by microbes. Furthermore, an interspecies an electron transfer mechanism in the coculture system was also reported where electron shuttle mediator, nanowires, and c-type cytochromes as membranebound proteins mediate to transfer an electron from one species to another. Methanobacterium spp., Methanobrevibacter spp., and Methanothermobacter spp., Methanoclculus spp. are the most commonly reported electromethanogens on the cathode (Fu et al., 2015; Ishii et al., 2019; Aryal et al., 2018). The process involved two possible pathways as shown in the following equations (Butler and Lovley, 2016; Nakasugi et al., 2017):

(17.2)

(17.3)

(17.4)

In a BES system for biogas upgrading, the reactor configuration, electrode materials, microbial dynamics, and applied potential have a significant impact on methane productivity. Not only limited to the methane, but also higher carbon chain chemicals and fuels synthesis are reported in the BES system utilizing CO2 and a poised electrode (Aryal et al., 2016, 2017; Aryal and Kvist, 2018; Bajracharya et al., 2019). The selection of a suitable electromethanogen, process optimization, and methane yield are the major bottlenecks for upscaling of the technology. From laboratory findings, it has been found that the CH4 formation rates and overall biochemical reaction efficiencies are still not satisfactory for rapid commercialization of the technology and demand further research (Nakasugi et al., 2017). The process constraints, such as the electron transfer rate, targeted product synthesis, electrode material, and low efficiency demand further research and prototype development. Similarly, redeg the bioreactor by removing the membrane separation of the anode from the cathode will make the reactor costeffective and simple, nevertheless O2 generation might affect the performance of anaerobes (Butler and Lovley, 2016; Giddings et al., 2015). In the future, an electromethanogenesis (EM) combined PtG system is expected to be a prominent technology for commercial biogas upgrading.

17.4.4 Photosynthetic biogas upgrading system

Photosynthetic CO2 utilization by microalgae is another type of biological biogas upgrading technology. This is based on utilization of CO2 from biogas as a carbon source and UV rays as an energy source by microalgae (prokaryotic cyanobacteria and eukaryotic microalgae) for their growth and multiplication. This upgrading process primarily requires the solubilization of CO2 into the nutrient-rich aqueous solution for the removal by microalgal photosynthesis (Posadas et al., 2015; Rodero et al., 2018), where CO2 fixation occurs by microalgae, and sulfur-oxidizing bacteria convert H2S to elemental sulfur and/or sulfate using O2 produced during the photosynthetic process or to sulfur dioxide by dissolved oxygen in an aqueous solution (Sun et al., 2015; Toledo-Cervantes et al., 2016). The microalgal biomass produced during this upgrading process

can be used for various applications such as feedstock for anaerobic digestion, biofuel generation, and bioactive compound extraction. Similarly, no inhibitory effect of CH4 was observed in microalgal biomass formation in the photobioreactor (Meier et al., 2015). The successful upgrading process can be designed using a two-stage process involving a photobioreactor connected with gas/liquid mass transfer equipment to inhibit oxygen desorption into the upgraded biogas, as well as monitoring of the liquid-to-biogas ratio (L/B) (Meier et al., 2015; Toledo-Cervantes et al., 2016). It is also important to optimize certain parameters like light intensity, pH, temperature, dissolved oxygen concentration, etc. to achieve high CO2 removal efficiency (Marín et al., 2018; Meier et al., 2019; Posadas et al., 2016). Different types of microalgal species are involved in biogas upgrading process, such as Arthrospira spp., Chlorella vulgaris, Chlorella kessleri, Cyanothece spp., Nannochloropsis gaditana, Limnothrix planktonica, and Phormidium spp. (Alcántara et al., 2015; Rodero et al., 2019; Toledo-Cervantes et al., 2016). Most of the photobioreactors used for biogas upgrading are of the enclosed type and are limited to the laboratory-scale only. Outdoor or open type of photobioreactors are simple in design and costeffective but have at higher risks of predator and microbial contamination. Toledo-Cervantes et al. (2016) found that the removal of CO2 and H2S from biogas was more than 99% effective, and of the desired quality to be injected directly into the natural gas grid. Similarly Meier et al. (2019) found that more than 80% of the CO2 from a biogas mixture is converted to biomass in 0.12 m³d −1 per m³ reactor volume. Likewise, a first demonstration-scale (proof-ofconcept) of photosynthetic biogas upgrading in an algal-bacterial photobioreactor showed that CH4 enrichment efficiency was ~90%, limited by N2 and O2 desorption (Rodero et al., 2019). From the different laboratory results, it can be concluded that this upgrading system is feasible at a commercial scale. However, not only the identification of key operational parameters is required but also the maximum treatment capacity of the process is determined in order to implement the system on a commercial scale.

17.5 Conclusion and future perspective

The different physicochemical biogas upgrading processes require high energy input and chemicals and do not utilize CO2 separated from the biogas mixture, which makes these technologies unsustainable. Hence, the alterative biological upgrading methods, which utilize CO2 from the biogas reactor for CH4 production as in PtG methanation and algal biomass generation as in photosynthetic upgrading methods have gained increasing attention in recent years. The existing biological upgrading technologies are not fully matured and optimized yet, and therefore needs further research and development. Hydrogenmediated microbial biogas upgrading has been viewed as a sustainable upgrading technology that utilizes CO2 to produce biomethane. The in situ microbial biogas upgrading technique has the limitation of H2 mass-liquid transfer and microbial dynamics shift that lead to the accumulation of volatile fatty acids, thereby inhibiting the methanogens. Furthermore, the unused molecular hydrogen during the upgrading process will also affect the Wobbe index of the upgraded biogas. Therefore it is important to maintain the right stoichiometric ratio of CO2:H2 which is 1:4; otherwise a high concentration of CO2 and/or H2 might affect the partial pressure, pH, and limit mass transfer that favors the termination of the methanation process, particularly in ex situ methanation. To resolve these problems, the hybrid ex situ and in situ technology could offer efficient conversion of CO2 to methane by overcoming the bottlenecks of in situ or ex situ methanation. Additionally, pure cultures of methanogens, like Methanothermobacter thermautotrophicus or Methanothermobacter marburgensis have been reported as efficient methanogens for CO2 conversion in ex situ methanation. Moreover, it may also be important to consider the gas recirculation flow rate and the reactor design to maximize the gas retention time and H2 dissolution to the liquid media. Similarly, the laboratory-scale photobioreactors need further optimization for their feasibility in large-scale production. Currently, biomethane production and utilization, both as transportation fuel and

gas grid injection, are dominated by European countries. However, the successful establishment of large-scale biomethane production and use in the global renewable energy market demands effective plans and policies from local, national, and international agencies, subsidies, technological advancements, market generation, sustainable biomass production, and cost effectiveness. In the future, large-scale biological biogas upgrading technologies will dominate the global market because there is the possibility of using genetically modified microbes for the effective treatment of various biomass upgrading processes in economically viable and environmentally friendly ways.

References

Adnan et al., 2019 Adnan AI, Ong MY, Nomanbhay S, Chew KW, Show PL. Technologies for biogas upgrading to biomethane: a review. Bioengineering. 2019;6(4):92 https://doi.org/10.3390/bioengineering6040092. Alcántara et al., 2015 Alcántara C, García-Encina PA, Muñoz R. Evaluation of the simultaneous biogas upgrading and treatment of centrates in a high-rate algal pond through C, N and P mass balances. Water Sci Technol. 2015;72(2):150–157 https://doi.org/10.2166/wst.2015.198. Angelidaki et al., 2018 Angelidaki I, Treu L, Tsapekos P, et al. Biogas upgrading and utilization: current status and perspectives. Biotechnol Adv. 2018;36(2):452–466 https://doi.org/10.1016/j.biotechadv.2018.01.011. Aryal and Kvist, 2018 Aryal N, Kvist T. Alternative of biogas injection into the Danish Gas Grid System—a study from demand perspective. ChemEngineering. 2018;2(3):43 https://doi.org/10.3390/chemengineering2030043. Aryal et al., 2016 Aryal N, Halder A, Tremblay PL, Chi Q, Zhang T. Enhanced microbial electrosynthesis with three-dimensional graphene functionalized cathodes fabricated via solvothermal synthesis. Electrochim Acta. 2016;217:117–122 https://doi.org/10.1016/j.electacta.2016.09.063. Aryal et al., 2017 Aryal N, Halder A, Zhang M, et al. Freestanding and flexible graphene papers as bioelectrochemical cathode for selective and efficient CO2 conversion. Sci Rep. 2017;7(1):1–8 https://doi.org/10.1038/s41598-017-098417. Aryal et al., 2018 Aryal N, Kvist T, Ammam F, Pant D, Ottosen LDM. An overview of microbial biogas enrichment. Bioresour Technol. 2018;264:359–369 https://doi.org/10.1016/j.biortech.2018.06.013. Awe et al., 2017 Awe OW, Zhao Y, Nzihou A, Minh DP, Lyczko N. A review of biogas utilisation, purification and upgrading technologies. Waste Biomass

Valoriz. 2017;8:267–283 https://doi.org/10.1007/s12649-016-9826-4. Bailera et al., 2017 Bailera M, Lisbona P, Romeo LM, Espatolero S. Power to gas projects review: lab, pilot and demo plants for storing renewable energy and CO2. Renew Sustain Energy Rev. 2017;69:292–312 https://doi.org/10.1016/j.rser.2016.11.130. Bailón and Hinge, 2012 Bailón, L., Hinge, J., 2012. Biogas and Bio-Syngas Upgrading—A Report. Danish Technological Institute. . Bajracharya et al., 2019 Bajracharya, S., Aryal, N., Wever, H., De Pant, D., 2019. Bioelectrochemical syntheses. In: Aresta, M., Karimi, I., Kawi, S. (Eds.), An Economy Based on Carbon Dioxide and Water. Springer, Cham, pp. 327– 358. . Bassani et al., 2015 Bassani I, Kougias PG, Treu L, Angelidaki I. Biogas upgrading via hydrogenotrophic methanogenesis in two-stage continuous stirred tank reactors at mesophilic and thermophilic conditions. Environ Sci Technol. 2015;49(20):12585–12593 https://doi.org/10.1021/acs.est.5b03451. Bassani et al., 2017 Bassani I, Kougias PG, Treu L, Porté H, Campanaro S, Angelidaki I. Optimization of hydrogen dispersion in thermophilic up-flow reactors for ex situ biogas upgrading. Bioresour Technol. 2017;234:310–319 https://doi.org/10.1016/j.biortech.2017.03.055. Bauer et al., 2013 Bauer F, Persson T, Hulteberg C, Tamm D. Biogas upgrading —technology overview, comparison and perspectives for the future. Biofuels Bioprod Bioref. 2013; https://doi.org/10.1002/bbb.1423. Benjaminsson et al., 2013 Benjaminsson, G., Benjaminsson, J., Rudberg, R.B., 2013. Power-to-gas – a technical review. In: SGC Rapport 2013:284. SGC, p. 67. . Blanco and Faaij, 2018 Blanco H, Faaij A. A review at the role of storage in energy systems with a focus on power to gas and long-term storage. Renew Sustain Energy Rev. 2018;81(1):1049–1086 https://doi.org/10.1016/j.rser.2017.07.062. Butler and Lovley, 2016 Butler CS, Lovley DR. How to sustainably feed a

microbe: strategies for biological production of carbon-based commodities with renewable electricity. Front Microbiol. 2016;7:1879 https://doi.org/10.3389/fmicb.2016.01879. Collet et al., 2017 Collet P, Flottes E, Favre A, et al. Techno-economic and life cycle assessment of methane production via biogas upgrading and power to gas technology. Appl Energy. 2017;192:282–295 https://doi.org/10.1016/j.apenergy.2016.08.181. Demirel and Scherer, 2011 Demirel B, Scherer P. Trace element requirements of agricultural biogas digesters during biological conversion of renewable biomass to methane. Biomass Bioenergy. 2011;35(2):992–998 https://doi.org/10.1016/j.biombioe.2010.12.022. Díaz et al., 2015 Díaz I, Pérez C, Alfaro N, Fdz-Polanco F. A feasibility study on the bioconversion of CO2 and H2 to biomethane by gas sparging through polymeric membranes. Bioresour Technol. 2015;185:246–253 https://doi.org/10.1016/j.biortech.2015.02.114. EBA, 2018 EBA, 2018. European Biogas Association Statistical Report. p. 15. . Electrochaea, 2018 Electrochaea, 2018. Electrochaea data sheet. <www.electrochaea.com>. Electrochaea, 2019 Electrochaea, 2019. Electrochaea fact sheet. <www.electrochaea.com>. Eveloy and Gebreegziabher, 2018 Eveloy V, Gebreegziabher T. A review of projected power-to-gas deployment scenarios. Energies. 2018;11(7):1824 https://doi.org/10.3390/en11071824. ForskEL, 2015 ForskEL, 2015. MeGa-StoRE – a final report. In: Energinet.dk Project No 12006. <www.lemvigbiogas.com>. ForskEL, 2016 ForskEL, 2016. MeGa-StoRE 2 phase 1—a final report. In: ForskEL 12270. <www.energiforskning.dk>. ForskNG, 2011 ForskNG, 2011. Main final report part 1—biogas–SOEC: Electrochemical upgrading of biogas to pipeline quality by means of SOEC

electrolysis. In: Project No. 10677. Frontera et al., 2017 Frontera P, Macario A, Ferraro M, Antonucci PL. ed catalysts for CO2 methanation: a review. Catalysts. 2017;7(2):59 https://doi.org/10.3390/catal7020059. Fu et al., 2015 Fu Q, Kuramochi Y, Fukushima N, Maeda H, Sato K, Kobayashi H. Bioelectrochemical analyses of the development of a thermophilic biocathode catalyzing electromethanogenesis. Environ Sci Technol. 2015;49(2):1225–1232 https://doi.org/10.1021/es5052233. Gahleitner, 2013 Gahleitner G. Hydrogen from renewable electricity: an international review of power-to-gas pilot plants for stationary applications. Int J Hydrogen Energy. 2013;38(5):2039–2061 https://doi.org/10.1016/j.ijhydene.2012.12.010. Gao et al., 2015 Gao J, Liu Q, Gu F, Liu B, Zhong Z, Su F. Recent advances in methanation catalysts for the production of synthetic natural gas. RSC Adv. 2015;5:22759–22776 https://doi.org/10.1039/c4ra16114a. Giddings et al., 2015 Giddings CGS, Nevin KP, Woodward T. Simplifying microbial electrosynthesis reactor design. Front Microbiol. 2015;6:1–6 https://doi.org/10.3389/fmicb.2015.00468. Gotz et al., 2016 Gotz M, Lefebvre J, Mors F, et al. Renewable power-to-gas: a technological and economic review. Renew Energy. 2016;85:1371–1390 https://doi.org/10.1016/j.renene.2015.07.066. Heide et al., 2010 Heide D, von Bremen L, Greiner M, Hoffmann C, Speckmann M, Bofinger S. Seasonal optimal mix of wind and solar power in a future, highly renewable Europe. Renew Energy. 2010;35(11):2483–2489 https://doi.org/10.1016/j.renene.2010.03.012. Høegh Jensen, 2015 Høegh Jensen, M., 2015. MeGa-stoRE. . Hoyer et al., 2016 Hoyer K, Hulteberg C, Svensson M, Nørregård, Ø. JJ. Biogas Upgrading—Technical Review. 76p Stockholm: Energiforsk; 2016. IEA Bioenergy, 2017 IEA Bioenergy, 2017. IEA Bioenergy Task 37: Country

Report Summaries. ISBN: 987-1-910154-34-2. IEA Bioenergy, 2018 IEA Bioenergy, 2018. Biogas in society—biological methanation demonstration plant in Allendorf, . An upgrading facility for biogas. In: IEA Bioenergy: Tasks 37. ‘Energy from biomass’. . Ishii et al., 2019 Ishii S, Imachi H, Kawano K, et al. Bioelectrochemical stimulation of electromethanogenesis at a seawater-based subsurface aquifer in a natural gas field. Front Energy Res. 2019;6:144 https://doi.org/10.3389/fenrg.2018.00144. Jentsch et al., 2014 Jentsch M, Trost T, Sterner M. Optimal use of power-to-gas energy storage systems in an 85% renewable energy scenario. Energy Procedia. 2014;46:254–261 https://doi.org/10.1016/j.egypro.2014.01.180. Joo et al., 2018 Joo SH, Delicio L, Muniz J, Baek S. Perspective: catalytic increase of biogas production in an anaerobic co-digestion system. Int J Nanopart Nanotechnol. 2018;4(1):1–6 https://doi.org/10.35840/2631-5084/5516. Kötter et al., 2015 Kötter E, Schneider L, Sehnke F, Ohnmeiss K, Schröer R. Sensitivities of power-to-gas within an optimised energy system. Energy Procedia. 2015;73:190–199 https://doi.org/10.1016/j.egypro.2015.07.670. Kougias and Angelidaki, 2018 Kougias PG, Angelidaki I. Biogas and its opportunities—a review. Front Environ Sci Eng. 2018;12:14 https://doi.org/10.1007/s11783-018-1037-8. Kvist and Aryal, 2019 Kvist T, Aryal N. Methane loss from commercially operating biogas upgrading plants. Waste Manage. 2019;87:295–300 https://doi.org/10.1016/j.wasman.2019.02.023. Lecker et al., 2017 Lecker B, Illi L, Lemmer A, Oechsner H. Biological hydrogen methanation – a review. Bioresour Technol. 2017;245:1220–1228 https://doi.org/10.1016/j.biortech.2017.08.176. Luo and Angelidaki, 2012 Luo G, Angelidaki I. Integrated biogas upgrading and hydrogen utilization in an anaerobic reactor containing enriched hydrogenotrophic methanogenic culture. Biotechnol Bioeng. 2012;109(11):2729–2736 https://doi.org/10.1002/bit.24557.

Luo et al., 2012 Luo G, Johansson S, Boe K, Xie L, Zhou Q, Angelidaki I. Simultaneous hydrogen utilization and in situ biogas upgrading in an anaerobic reactor. Biotechnol Bioeng. 2012;109(4):1088–1094 https://doi.org/10.1002/bit.24360. Luo et al., 2014 Luo G, Wang W, Angelidaki I. A new degassing membrane coupled upflow anaerobic sludge blanket (UASB) reactor to achieve in situ biogas upgrading and recovery of dissolved CH4 from the anaerobic effluent. Appl Energy. 2014;132:536–542 https://doi.org/10.1016/j.apenergy.2014.07.059. Marín et al., 2018 Marín D, Posadas E, Cano P, et al. Seasonal variation of biogas upgrading coupled with digestate treatment in an outdoors pilot scale algal-bacterial photobioreactor. Bioresour Technol. 2018;263:58–66 https://doi.org/10.1016/j.biortech.2018.04.117. Meier et al., 2015 Meier L, Pérez R, Azócar L, Rivas M, Jeison D. Photosynthetic CO2 uptake by microalgae: an attractive tool for biogas upgrading. Biomass Bioenergy. 2015;73:102–109 https://doi.org/10.1016/j.biombioe.2014.10.032. Meier et al., 2019 Meier L, Martínez C, Vílchez C, Bernard O, Jeison D. Evaluation of the feasibility of photosynthetic biogas upgrading: simulation of a large-scale system. Energy. 2019;189:116313 https://doi.org/10.1016/j.energy.2019.116313. Mendoza-Hernandez et al., 2019 Mendoza-Hernandez OS, Shima A, Matsumoto H, et al. Exergy valorization of a water electrolyzer and CO2 hydrogenation tandem system for hydrogen and methane production. Sci Rep. 2019;9:6470 https://doi.org/10.1038/s41598-019-42814-6. Miltner et al., 2016 Miltner M, Makaruk A, Harasek M. Selected methods of advanced biogas upgrading. Chem Eng Trans. 2016;52:463–468 https://doi.org/10.3303/CET1652078. Miltner et al., 2017 Miltner M, Makaruk A, Harasek M. Review on available biogas upgrading technologies and innovations towards advanced solutions. J Cleaner Prod. 2017;161:1329–1337 https://doi.org/10.1016/j.jclepro.2017.06.045. Mshandete et al., 2006 Mshandete A, Björnsson L, Kivaisi AK, Rubindamayugi

MST, Mattiasson B. Effect of particle size on biogas yield from sisal fibre waste. Renew Energy. 2006;31(14):2385–2392 https://doi.org/10.1016/j.renene.2005.10.015. Muñoz et al., 2015 Muñoz R, Meier L, Diaz I, Jeison D. A review on the stateof-the-art of physical/chemical and biological technologies for biogas upgrading. Rev Environ Sci Biotechnol. 2015;14(4):727–759 https://doi.org/10.1007/s11157-015-9379-1. Nakasugi et al., 2017 Nakasugi Y, Himeno M, Kobayashi H, Ikarashi M, Maeda H, Sato K. Experimental and mathematical analyses of bio-electrochemical conversion of carbon dioxide to methane. Energy Procedia. 2017;114:7133–7140 https://doi.org/10.1016/j.egypro.2017.03.1857. Niaz et al., 2015 Niaz S, Manzoor T, Pandith AH. Hydrogen storage: materials, methods and perspectives. Renew Sustain Energy Rev. 2015;50:457–469 https://doi.org/10.1016/j.rser.2015.05.011. Niesner et al., 2013 Niesner J, Jecha D, Stehlík P. Biogas upgrading technologies: state of art review in European region. Chem Eng Trans. 2013;35:517–522 https://doi.org/10.3303/CET1335086. Perez-Sanz et al., 2019 Perez-Sanz FJ, Sarge SM, van der Veen A, Culleton L, Beaumont O, Haloua F. First experimental comparison of calorific value measurements of real biogas with reference and field calorimeters subjected to different standard methods. Int J Therm Sci. 2019;135:72–82 https://doi.org/10.1016/j.ijthermalsci.2018.06.034. Posadas et al., 2015 Posadas E, Serejo ML, Blanco S, Pérez R, García-Encina PA, Muñoz R. Minimization of biomethane oxygen concentration during biogas upgrading in algal-bacterial photobioreactors. Algal Res. 2015;12:221–229 https://doi.org/10.1016/j.algal.2015.09.002. Posadas et al., 2016 Posadas E, Szpak D, Lombó F, et al. Feasibility study of biogas upgrading coupled with nutrient removal from anaerobic effluents using microalgae-based processes. J Appl Phycol. 2016;28:2147–2157 https://doi.org/10.1007/s10811-015-0758-3. Prussi et al., 2019 Prussi M, Padella M, Conton M, Postma ED, Lonza L. Review of technologies for biomethane production and assessment of Eu

transport share in 2030. J Cleaner Prod. 2019;222:565–572 https://doi.org/10.1016/j.jclepro.2019.02.271. Rachbauer et al., 2016 Rachbauer L, Voitl G, Bochmann G, Fuchs W. Biological biogas upgrading capacity of a hydrogenotrophic community in a trickle-bed reactor. Appl Energy. 2016;180:483–490 https://doi.org/10.1016/j.apenergy.2016.07.109. Rachbauer et al., 2017 Rachbauer L, Beyer R, Bochmann G, Fuchs W. Characteristics of adapted hydrogenotrophic community during biomethanation. Sci Total Environ. 2017;595:912–919 https://doi.org/10.1016/j.scitotenv.2017.03.074. Rasmussen and Aryal, 2019 Rasmussen NBK, Aryal N. Syngas production using straw pellet gasification in fluidized bed allothermal reactor under different temperature conditions. Fuel. 2019;263:116706 https://doi.org/10.1016/j.fuel.2019.116706. Rodero et al., 2018 Rodero MR, Ángeles R, Marín D, et al. Biogas Purification and Upgrading Technologies Cham: Springer; 2018;239–276 http://doi.org/10.1007/978-3-319-77335-3_10. Rodero et al., 2019 Rodero M, del R, Lebrero R, et al. Technology validation of photosynthetic biogas upgrading in a semi-industrial scale algal-bacterial photobioreactor. Bioresour Technol. 2019;279:43–49 https://doi.org/10.1016/j.biortech.2019.01.110. Ryckebosch et al., 2011 Ryckebosch E, Drouillon M, Vervaeren H. Techniques for transformation of biogas to biomethane. Biomass Bioenergy. 2011;35(5):1633–1645 https://doi.org/10.1016/j.biombioe.2011.02.033. Schaaf et al., 2014 Schaaf T, Grünig J, Schuster MR, Rothenfluh T, Orth A. Methanation of CO2—storage of renewable energy in a gas distribution system. Energy Sustain Soc. 2014;4 https://doi.org/10.1186/s13705-014-0029-1. Scholz et al., 2013 Scholz M, Melin T, Wessling M. Transforming biogas into biomethane using membrane technology. Renew Sustain Energy Rev. 2013;17:199–212 https://doi.org/10.1016/j.rser.2012.08.009. Scholz et al., 2015 Scholz M, Alders M, Lölsberg J, Wessling M. Dynamic

process simulation and process control of biogas permeation processes. J Membr Sci. 2015;484:107–118 https://doi.org/10.1016/j.memsci.2015.03.008. Schwede et al., 2017 Schwede S, Bruchmann F, Thorin E, Gerber M. Biological syngas methanation via immobilized methanogenic archaea on biochar. Energy Procedia. 2017;105:823–829 https://doi.org/10.1016/j.egypro.2017.03.396. Seifert et al., 2014 Seifert AH, Rittmann S, Herwig C. Analysis of process related factors to increase volumetric productivity and quality of biomethane with Methanothermobacter marburgensis. Appl Energy. 2014;132:155–162 https://doi.org/10.1016/j.apenergy.2014.07.002. Strevett et al., 1995 Strevett KA, Vieth RF, Grasso D. Chemo-autotrophic biogas purification for methane enrichment: mechanism and kinetics. Chem Eng J Biochem Eng J. 1995;58(1):71–79 https://doi.org/10.1016/0923-0467(95)060952. Strübing et al., 2017 Strübing D, Huber B, Lebuhn M, Drewes JE, Koch K. High performance biological methanation in a thermophilic anaerobic trickle bed reactor. Bioresour Technol. 2017;245:1176–1183 https://doi.org/10.1016/j.biortech.2017.08.088. Sun et al., 2015 Sun Q, Li H, Yan J, Liu L, Yu Z, Yu X. Selection of appropriate biogas upgrading technology-a review of biogas cleaning, upgrading and utilisation. Renew Sustain Energy Rev. 2015;51:521–532 https://doi.org/10.1016/j.rser.2015.06.029. Sveinbjörnsson et al., 2017 Sveinbjörnsson, D., Münster, E., Aryal, N., Bo, R., Pedersen, B., 2017. WP1 Gas Conditioning and Grid Operation Deliverable 1.1.1 Upgrading of Biogas to Biomethane With the Addition of Hydrogen from Electrolysis. <www.futuregas.dk>. Thema et al., 2019a Thema M, Bauer F, Sterner M. Power-to-gas: electrolysis and methanation status review. Renew Sustain Energy Rev. 2019a;112:775–787 https://doi.org/10.1016/j.rser.2019.06.030. Thema et al., 2019b Thema M, Weidlich T, Hörl M, et al. Biological CO2methanation: an approach to standardization. Energies. 2019b;12(9):1670 https://doi.org/10.3390/en12091670.

Thrän UFZ et al., 2014 Thrän UFZ, D., Billig Tobias Persson, E., Svensson, M., Daniel-Gromke, J., Ponitka, J., Seiffert John Baldwin, M., et al., 2014. Biomethane-status and factors affecting market development and trade. In: IEA Bioenergy, Task 40 and Task 37. . Toledo-Cervantes et al., 2016 Toledo-Cervantes A, Serejo ML, Blanco S, Pérez R, Lebrero R, Muñoz R. Photosynthetic biogas upgrading to bio-methane: boosting nutrient recovery via biomass productivity control. Algal Res. 2016;17:46–52 https://doi.org/10.1016/j.algal.2016.04.017. Treu et al., 2018 Treu L, Campanaro S, Kougias PG, Sartori C, Bassani I, Angelidaki I. Hydrogen-fueled microbial pathways in biogas upgrading systems revealed by genome-centric metagenomics. Front Microbiol. 2018;9:1079 https://doi.org/10.3389/fmicb.2018.01079. Ullrich et al., 2018 Ullrich T, Lindner J, Bär K, Mörs F, Graf F, Lemmer A. Influence of operating pressure on the biological hydrogen methanation in trickle-bed reactors. Bioresour Technol. 2018;247:7–13 https://doi.org/10.1016/j.biortech.2017.09.069. Vo et al., 2018 Vo TTQ, Wall DM, Ring D, Rajendran K, Murphy JD. Technoeconomic analysis of biogas upgrading via amine scrubber, carbon capture and ex-situ methanation. Appl Energy. 2018;212:1191–1202 https://doi.org/10.1016/j.apenergy.2017.12.099. Voelklein et al., 2019 Voelklein MA, Rusmanis D, Murphy JD. Biological methanation: strategies for in-situ and ex-situ upgrading in anaerobic digestion. Appl Energy. 2019;235:1061–1071 https://doi.org/10.1016/j.apenergy.2018.11.006. Wahid et al., 2019 Wahid R, Mulat DG, Gaby JC, Horn SJ. Effects of H2:CO2 ratio and H2 supply fluctuation on methane content and microbial community composition during in-situ biological biogas upgrading. Biotechnol Biofuels. 2019;12:104 https://doi.org/10.1186/s13068-019-1443-6. Weiland, 2010 Weiland P. Biogas production: current state and perspectives. Appl Microbiol Biotechnol. 2010;85(4):849–860 https://doi.org/10.1007/s00253009-2246-7.

Wulf et al., 2018 Wulf C, Linßen J, Zapp P. Review of power-to-gas projects in Europe. Energy Procedia. 2018;155:367–378 https://doi.org/10.1016/j.egypro.2018.11.041. Xia et al., 2016 Xia A, Cheng J, Murphy JD. Innovation in biological production and upgrading of methane and hydrogen for use as gaseous transport biofuel. Biotechnol Adv. 2016;34(5):451–472 https://doi.org/10.1016/j.biotechadv.2015.12.009. Yousef et al., 2016 Yousef AMI, Eldrainy YA, El-Maghlany WM, Attia A. Upgrading biogas by a low-temperature CO2 removal technique. Alex Eng J. 2016;55(2):1143–1150 https://doi.org/10.1016/j.aej.2016.03.026. Zabranska and Pokorna, 2018 Zabranska J, Pokorna D. Bioconversion of carbon dioxide to methane using hydrogen and hydrogenotrophic methanogens. Biotechnol Adv. 2018;36(3):707–720 https://doi.org/10.1016/j.biotechadv.2017.12.003.

Index

Note: Page numbers followed by “f” and “t” refer to figures and tables, respectively.

A

Absorption, 74–75

absorption-combinations, 166

biogas upgrading by, 145–148

columns, 35–42

configurations, 62–64

hollow fiber membrane ors, 63–64

packed column reactors, 62–63

methodologies, 58–62, 276–277

in K2CO3 solutions, 61–62

liquid/gas ratios for water scrubbing purposes, 59t

in NaOH solutions, 60–61

in water, 58–60

Acetic acid, 16–17, 209, 333, 414

Acetobacterium woodii, 348–349

Acetoclastic methanogens, 327

Acetogenesis, 9, 10, 414

Acetyl-co-enzyme A, 10

Acid gases, 29

Acidification, 427

Acidogenesis, 9, 9–10, 414

Acidogenic bacteria, 10–11

Acidogenic microorganisms, 328–329

Aciplex, 194

Actinobacillus succinogenes, 421

Actinobacteria, 12–13

Activated carbon (AC), 104, 259

adsorption capacity of, 262–264

adsorption into, 260–266

adsorption onto, 300–301

mutual displacement of volatile methylsiloxanes from, 264

possibilities of activated carbon regeneration, 265–266

preparation of biogas for adsorption into, 264

and regeneration of silica gel, 267–268

Activated MDEA (aMDEA), 31, 33–34, 47

Activation energy (Ea), 191–192

Adiabatic reactors, 198–199, 199

Adsorbents in PSA, 103–107

Adsorption, 95, 98

into activated carbon, 260–266

adsorption-combinations, 166

into alumina, 271–274

capacity

of activated carbon, 262–264

and regeneration of activated Al2O3, 272–274

and regeneration of novel polymer adsorbents, 275–276

and regeneration of zeolites, 269–271

isotherm of PSA, 107–109, 108t

kinetics of PSA, 109–112

Knudsen diffusivity, 110

molecular diffusion, 110

Poiseuille diffusion or viscous diffusion, 110–112

methods, 260–276

into polymer adsorbents, 274–276

preparation of biogas for adsorption into activated carbon, 264

into silica gel, 266–268

time, 101–102

into zeolites, 268–271

Advection, 339

Aerobic biological air filters, 309–310

Afterburners, 67

Agriculture policies and regulations, 454

Agrobacterium, 280

Airlift PBRs, 391–392

Alkaline electrolysis, 324

Alkaline electrolyzers, 190–193

developments, 193

reactor configurations, 191–192

Alkaline pH microalgal system, 394

Alkaline scrubbing, 60–61

Alkaline solid polymeric membrane electrolyzer (Alkaline-SPE), 474–476

Alkaline water electrolysis (AWE), 189–190, 190–191, 191–192

Alkaline with regeneration (AwR), 425

Allocation, 422

Alumina, adsorption into, 271–274

Ambient temperature, 12

Amine absorption/stripping, 29

for biogas upgrading, 30

economic considerations, 45–46

operational problems and emissions, 43–45

process description and technology, 34–42

process fundamentals, 31–46

research and development directions, 47–50

Amine chemistry, 31–32

Amine concentration, 65

Amine losses, 43

Amine scrubbing, 29, 47

Amine selection, 32–34

Amine solutions, 224

Amine solvents, 31

Amine-functionalized solid sorbents, 48–49

Amino acid solutions (AASs), 47

2-Amino-2-methyl-1-propanol (AMP), 32–33

Ammonia (NH3), 65, 167–168, 223–224, 337, 363–364, 413–414, 415–416, 416–417, 467–468, 468

Anaerobic digesters, 328

Anaerobic digestion (AD), 8–14, 363, 383–384, 413–414, 449–452

mechanism, 9–11

processes in, 328–330

successful policies in anaerobic digestion implementation, 452–460

policies and regulations, 452

Anion exchange membranes (AEM), 231

Anoxic biological air filters, 308–309

APTMDS, See Bis(3-aminopropyl)-tetramethyldisiloxane (APTMDS)

Arthrobacter, 279, 280

Arthrospira spp., 484

A. platensis, 394

Atmospheric pressure, 30

Attributional LCA (ALCA), 421

environmental performance of biomethane, 428

of hydrate-based biogas upgrading technology, 426

Audi e-gas technology, 476–478

B

Bacteroidetes, 12–13

Barrier model, 111

Bi-LDF model, 111

Bio liquefied natural gas (bio-LNG), 162

Bio-CNG, 74

Biocathodes, 209

Bioelectrochemical biogas upgrading, fundamentals of, 364–368

Bioelectrochemical methanation, 189, 205–211 See also Electrocatalytic methanation

biocathodes, 209

direct electron transfer, 207–208


Bioelectrochemical systems (BES), 7, 205–206, 210f, 212, 364, 468, 482–484

economical insights, 375–376

fundamentals of bioelectrochemical biogas upgrading, 364–368

methane enrichment of biogas, 368–374

prospective and challenges, 376–378

Biofilms, 209

Biofilters (BFs), 278–279, 306

Biofuels Quota Act of 2007, 456

Biogas, 73–74, 159–160, 223–224, 255, 363–364, 383–384, 413–414, 467–468

anaerobic digestion, 8–14

biohythane, 14–16

challenges and way forward, 18–19

cleaning, 384–385

CO2 removal approaches from biogas, 74–75

combined methods for volatile organic silicon compounds removal from biogas, 281

comparison of methods for reducing the content of volatile organic silicon compounds in, 281–285

composition, 395

and feedstock types, 468

constituents, 469t

electrochemically induced biogas upgradation, 16–18

factors affecting biogas production, 11–14

methods for reducing content of volatile organic silicon compounds in, 258–281

plants, 414

production, 118–119

purification, 311

state-of-the-art of biogas production, 4–6

trends in biogas utilization, 6–8

utilization, 451–452

Biogas Action Program, 452

Biogas upgrading/upgradation (BU), 3–4, 6f, 74, 159–160, 298, 306, 321–323, 384–385 See also Photosynthetic biogas upgrading

by absorption and hybrid absorption-membrane, 145–148

advantages and disadvantages, 419t

approach and challenges, 241–246

biogas and scale-up approaches, 244–246

CO2 electroreduction, 241–243

H2S oxidation, 243–244

biomethanation, 415–421

by cryogenic and hybrid cryogenic-membrane separation, 143–145

life cycle assessment, 421–423

membrane configurations and plant design for, 140–142

for natural gas grid injection and transport fuel, 468–471

state-of-the-art of, 4–6

technologies, 166

Biogasclean QSR and MUW process, 310

Biohydrogen, 14

Biohythane, 14–16

Biological air filtration, 306–311, 307f

aerobic biological air filters, 309–310

anoxic biological air filters, 308–309

commercial applications, 310–311

Biological gas cleaning, 278–279

Biological mediated methanation, 212

Biological methanation technology, 326–327

Biological methods, 278–281

of BU, 5

Biological reactors, 224–225

Biological upgrading

biogas upgrading, 321–323

factors controlling biomethanation, 337–350

hydrogen generation and utilization, 324–326

methanation, 326–327

microbial basis for biomethanation, 327–330


standards for natural gas quality, 322t

Biomethanation, 415–421, 448–449 See also Catalytic/thermochemical methanation, Chemoautotrophic methanation

cleaning of biogas, 416–417

factors controlling, 337–350

bacterial interaction and competition, 348–350

elemental composition of cellular biomass, 345t

growth requirements, 344–346

mass transfer of H2, 337–343

in situ and ex situ biomethanation reactor performances, 334t

strengths, weaknesses, opportunities, threats analysis, 355t

temperature, 343–344

microbial basis for, 327–330

Gibbs free energy of acetogenic processes, 330t

methanogens, 327

processes in anaerobic digestion, 328–330

upgrading of biogas into biomethane, 417–421

Biomethane, 3–4, 57–58, 74, 159–160, 168, 298, 322, 384–385

comparison of biomethane standards, 471t

decision- system for biomethane implantation with techno-economic analysis and policies, 461–462

policy instruments, 456–460

successful policies in anaerobic digestion implementation, 452–460

agriculture policies and regulations, 454

incentives, 455

policies and regulations, 452

renewable energy-related policies and regulations, 453–454

waste management policies, 454

techno-economic studies in anaerobic digestion, 449–452

technologies, 448

upgrading of biogas into, 417–421

absorption, 417

adsorption, 417–418

cryogenic separation, 418–421

membrane separation, 418

Biopuric process, 310

Bioscrubbers (BSs), 278, 306

Biosynthetic natural gas (BioSNG), 322

Biotrickling filters (BTFs), 279, 306

Bis(3-aminopropyl)-tetramethyldisiloxane (APTMDS), 140

Blowdown time, 102

Borate, 65

Boric salts, 65

Bottom ash upgrading (BABIU), 425

Brunauer-Emmett-Teller (BET)

isotherm, 108

surface area, 104

Bubble columns, 35, 37–39

Bubble-cap tray, 37

Bubbling type reactor, 201–202

Buchanan equation, 41–42

Butyl-trimethylammonium bis(trifluoromethylsulfonyl) imide ((N1114) (BTA)), 202–203

C

Calorific value, 321–322

Cambrian Innovation, 482–483

Capital expenses (CAPEX), 343–344

Carbon

capture or recycling, 205–206

carbon-based adsorbents, 104–106, 105f

activated carbons, 104

carbon molecular sieves, 104–106

cloth, 209

felt, 209

membranes, 448–449

plates, 209

Carbon and nitrogen ratio (C/N ratio), 13

Carbon capture and storage technologies (CCS technologies), 188

Carbon dioxide (CO2), 4–5, 5, 32–33, 93–94, 223–224, 225, 321–322, 346–347, 363–364, 413–414, 467–468

biomethanation, 479

CO2R process, 232, 232

conversion in to succinic acid, 421

cryogenic condensation techniques, 172–173

electroreduction, 241–243

electroreduction of, 226–236

reduction to methane, 189

removal

approaches from biogas, 74–75

from biogas, 119

effects of operating parameters on CO2 removal in water scrubber, 78–80

efficiency, 75

through photosynthetic-bacterial associated biogas upgradation, 386–387

separation, 450–451

utilization, 420

Carbon molecular sieves (CMSs), 104, 104–106, 118

membranes, 132–134

Carbon monoxide (CO), 363–364

Carbonization, 132–133, 260

Carboxylic acids, 44

Catalyst setting (CS), 200

Catalytic/thermochemical methanation, 476–478, 477t

ETOGAS and Audi e-gas technology, 476–478

Haldor Topsøe, 478

methane gas storage of renewable energy, 478

Cathode materials, 209

Cathodic electron transfer, 368

Cation exchange membranes (CEM), 231

Cellulose acetate (CA), 117–118

Channeling, 81, 82

Chapman-Enskog equation, 110

Chemical absorption, 29, 29–30, 472–473

advantages and disadvantages, 30t

economic considerations, 45–46

method, 74–75

operational problems and emissions, 43–45

process

description and technology, 34–42

fundamentals, 31–46

research and development directions, 47–50

tests, 276

Chemical methanation, 189, 326

Chemical oxidation, 258

Chemical promoters in water absorption, 64–66

Chemical solvents, 224

Chemoautotrophic methanation, 478–482, 481t

bioelectrochemical system, 482–484

Electrochaea, 480–482

MicrobEnergy, 482

photosynthetic biogas upgrading system, 484–485

Chemotrophs, 309

Chlorella, 386, 386–387, 392, 394

C. kessleri, 484

C. sorokiniana-based biogas upgradation study, 394–395

C. vulgaris, 484

MM-2, 395

Chlorides (Cl−), 223–224

Chlorococcum, 386

Claus process, 224

Cleaning of biogas, 416–417

H2S removal, 416

impurities removal, 416–417

water removal, 416

Clogging, 86–87

Clostridium butyricum, 11–12

Cobaltites (LaCoO3), 197

Column length, 103

Combined heat and power plants (CHP), 467–468

Combustion, 296

COMFLUX process, 200–201

Commercial biogas upgrading systems, 4–5

Commercial organic solvents, 224

Composite membrane, 123

Compressed natural gas (CNG), 14–15

Condensation, 259

Consequential LCA (CLCA), 421

Container-based SYstem for MethAnation (COSYMA), 200–201, 201f

Continuous stirred tank reactors (CSTR), 333, 479

Conventional amine absorbents, 47

Cool energy supply, 173–174

Cooling, 259

Correlation factor, 208

Corrosion process, 296

Coulombic efficiency (CE), 366–367

Credits

for carbon reduction and carbon trading, 455

for nutrient load reduction, 455

for renewable energy and renewable transportation fuel, 455

Cryogenic distillation, 162–163, 162f, 164f

Cryogenic hybrid systems, 160

Cryogenic membrane separation, biogas upgrading by, 143–145

Cryogenic packed bed technology (B technology), 163–166

Cryogenic separation, 74–75

Cryogenic techniques, 160

comparison of documented technologies, 177–178

cryogenic biogas upgrading, 160–166

cryogenic hybrid systems, 166–171

full-scale experiences and technoeconomic studies, 176–177

potential combination of cryogenic and membrane processes, 170

Cryogenic-absorption combination process, 167–168

Cryogenic-adsorption synergized process, 168–169

Cryogenic-based biogas upgradation, 385

Cryogenic-hydrate processes, 170–171

Cryogenic-membrane processes, 160, 171–176

Cumulative energy demand (CED), 425

Cumulative nonrenewable energy demand, 427

Cyanothece spp., 484

D

D4 adsorption process, 269

Decision methodology (DSM), 461–462

Decision- for Digester-Algae IntegRation for Improved Environmental and Economic Sustainability (DAIRIEES), 461–462

Decision- system for biomethane implantation with techno-economic analysis and policies, 461–462

Deferribacteres, 12–13

Degradation of absorbent, 44

Demister, 86

Dense membranes, 122

Desorption, 98

columns, 35–42

Desulfurization, 300

Di-methyl ethanol amine, 74–75

Diameter of scrubbing column, 81–82

Diethanolamine (DEA), 31, 64–65, 417, 472–473

Diffusion, 340

through micropores, 111

Diglycolamine (DGA), 472–473

Dimethyldioxirane, 258

Dimethylsilanediol (DMSD), 279

Direct electron transfer (DET), 206, 207–208, 368

Direct interspecies electron transfer (DIET), 4

Dissolved inorganic carbon (DIC), 338

Divinylbenzene (DVB), 274–275

DMF, See N,N-dimethylformamide (DMF)

Dosing iron salts/oxides into digester, 313–314

Dry packing, 86

Dry reforming process, 427

Dual resistance model, 111

Dual-mode-sorption model (DMS model), 137, 138

Dunwald-Wagner model, 111

E

Economical insights, 375–376

Electric swing adsorption (ESA), 95, 97

Electricity consumption for water scrubbing, 66, 66t

Electrocatalytic methanation, 189–205 See also Bioelectrochemical methanation

alkaline electrolyzers, 190–193

fixed-bed methanation reactors, 198–200

fluidized bed methanation reactors, 200–202

microchannel reactors, 203–205

polymer electrolyte membrane electrolyzers, 193–196

solid oxide electrolyzers, 196–198

three-phase reactor, 202–203

Electrochaea, 342, 342–343, 480–482

Electrochemical approach for biogas upgrading

biogas upgrading approach and challenges, 241–246

electrochemical oxidation of H2S, 236–241

electroreduction of CO2, 226–236

Faradaic and energy efficiency, 226

Electrochemical oxidation of H2S, 236–241

considerations, 237

reactor and process design, 237–241

Electrochemical reduction, See Electroreduction of CO2

Electrochemically induced biogas upgradation, 16–18

Electrofermentation (EF), 8

Electrolyzers, 189–190

advantages and challenges, 242t

alkaline, 190–193

polymer electrolyte membrane, 193–196

solid oxide, 196–198

Electromethanogenesis (EM), 4, 368

Electromethanogens, 368

Electron donors, 3–4

Electron transfer mechanism, 368–370

Electroreduction of CO2, 226–236

considerations, 227–230

liquid electrolyte devices, 229–230

solid-oxide devices, 228–229

designs, 230f

operating conditions and product performances, 233t

reactor and process design, 230–236

Elemental sulfur, 387

Embden–Meyerhof–Parnas pathway, 9–10

Empty bed retention time (EBRT), 307

Energy

consumption, 42–43, 66–67

dynamics, 4

policy, 458

Energy efficiency (EE), 226, 367, 368

of methane upgrading, 66

Entner–Doudoroff pathway, 9–10

Entrainment eliminator, 86

2-Ethoxyethanol (2EE), 48

1-Ethyl-3-methylimidazolium trifluoromethanesulfonate ((EMIM) (Tf)), 202– 203

ETOGAS, 476–478

Eutrophication, 427

Ex situ


process, 479

removal using sulfur-oxidizing microorganisms, 306–312

biological air filtration, 306–311, 307f

microalgal removal of H2S, 311–312

External or film diffusion, 109

Extract product, 98

F

Faradaic efficiency, 226

Favorable isotherm curve, 109

Feed stocks, 450

Feed-in tariff (FIT), 455

Fermentation processes, 9

Ferric chloride (FeCl3), 313

Ferric hydroxide (Fe(OH)3), 313

Ferricyanide, 240

Ferrites (LaFeO3), 197

Ferrous sulfate, 314

Ferrous sulfide (FeS), 313

Fick’s first law, 136

First-generation amine scrubbing technology, 29

Fischer–Tropsch synthesis (FT synthesis), 427

5Å sieve, 106

Fixed-bed methanation reactors, 198–200

design, 198–199

developments, 199–200

reactor configurations, 199

Fixed-bed reactor (FBR), 476

Flat sheet, 124

Flat-plate PBRs, 391–392

Flemion, 194

Flooding point of packing, 41–42

Flow rate, 103

Fluidized bed methanation reactors, 200–202

design, 200



Foaming, 44–45

Free volume of polymer membrane, 130

Freezing methods, 259, 259–260

Freundlich isotherm, 108

Fuel cell concept, 237–239

Fumapem, 194

Fusarium oxysporum, 279, 280

G

GaBi software, 428–429

Gammaproteobacteria, 12–13

Gas diffusion electrodes (GDEs), 193, 232

Gas diffusion layer (GDL), 193

Gas engines, 321–322

Gas film flux, 341

Gas flow (GF), 199, 200

rate, 79–80, 395–396

Gas grid system, 468–471

Gas load factor, 41

Gas permeation unit (GPU), 121

Gas purification technology, 450–451

Gas separation

basic membrane processes for, 117–120, 118f

principal membrane companies and membrane materials, 129t

theory of transport in gas separation on membranes, 135–140

transport equations through glassy polymers, 137–140

transport through rubbery polymers, 135–137

Gas treatment services (GtS), 161

Gas–liquid absorption, 276

membranes for biogas upgrading, 127

Gas–liquid ors, 35, 35

Genosorb, 74–75

Geobacter, 208

G. metallireducens, 208

Glassy polymers, 127, 129–130

transport equations through, 137–140

Global warming, 427

Global warming potential (GWP), 414–415

Graham diffusion theory, 117–118

Graham law of diffusion, 122

Graphite, 209

Greenhouse gas emissions (GHG emissions), 383–384, 414–415

reduction targets, 454

H

H-type cell, 231

Halospirulina sp., 394

Height of packing, 40

Height of scrubbing column, 82–83

Height of transfer unit (HTU), 82–83

Henderson–Hasselbalch equation, 349

Henry’s constants, 59, 59t

Henry’s isotherm, 107

Henry’s law, 135, 337, 338t

Heyrovsky reaction, 190–191

High pressure water scrubbing (HPWS), 425

High rate algal pond (HRAP), 427

High-pressure water scrubbing, 450–451

High-rate algal ponds (HRAPs), 388–391

Hollow fibers (HF), 117–118, 124, 125f

membrane ors, 63–64

Homoacetogenesis, 329, 350

Homoacetogenic bacteria, 348

Homoacetogens, 16–17

Homogeneous solid diffusion model, 111

Hot potassium carbonate, 61–62

Hybrid absorption-membrane, biogas upgrading by, 145–148

Hybrid cryogenic-membrane separation, biogas upgrading by, 143–145

Hybrid technologies, 420

Hydrate-combinations, 166

Hydraulic retention time (HRT), 13–14

Hydrocarbons, 223–224

Hydrogen (H2), 5, 321–322, 363–364, 413–414

evolution, 369

gas–liquid mass transfer rate, 341–342

generation and utilization, 324–326

partial pressure in gas phase, 342

Hydrogen evolution reaction (HER), 194–195, 228

Hydrogen sulfide (H2S), 59–60, 223–224, 225, 236, 295–296, 321–322, 363– 364, 383–384, 413–414, 415–416, 467–468, 468

advantages and disadvantages of techniques for removal of, 315t

combined chemical-biological processes, 314

comparison of H2S removal techniques, 314–316

electrochemical oxidation, 236–241

ex situ removal using sulfur-oxidizing microorganisms, 306–312

oxidation, 243–244

physicochemical removal technologies, 298–306

removal, 416

through photosynthetic-bacterial associated biogas upgradation, 386–387

in situ H2S removal, 312–314

technologies for removal of biogas contaminants, 297–298

tolerance limits, 295–296

Hydrogenotrophic methanogenesis, 16, 212, 369

Hydrogenotrophic methanogens, 327, 479

Hydrogenotrophs, 10–11

Hydrolysis, 9, 9–10, 11, 414

Hydrolytic bacteria, 328–329

I

Ideal membrane selectivity, 121

Immobilized liquid membranes, 306

Impurities removal, 416–417

In situ


CH4 enrichment, 420

H2S removal, 312–314

dosing iron salts/oxides into digester, 313–314

in situ microaeration, 312–313

process, 479

Incentives, 455

Indirect electron transfer (IET), 206, 368

Industrial lung, 420

“Initiative for Natural-Gas-Based Mobility”, 456

Inorganic membranes for gas separation, 132

Interfacial area, 342

Internal or intraparticle diffusion, 109

Interpretation, 423

Ion-exchange membrane, 364

Ionic liquids, 202–203

Iron (III) salts, 313

Iron oxide pellets, 302

Iron oxide wood chips, 302

Isothermal reactors, 199

J

Joule effect, 97

K

Knudsen diffusivity, 110

Kryosol process, 277

L

Langmuir “hole-filling” process, 137

Langmuir isotherm, 108

Lanthanum manganites (LaMnO3), 197

Large-scale biogas upgrading plants

biogas composition and feedstock types, 468

biogas upgrading for natural gas grid injection and transport fuel, 468–471

comparison of biomethane standards, 471t

state-of-the-art of large-scale biogas upgrading technologies, 472–485

LaxSr1–xVO3–δ (LSV), 240–241

Leptolyngbya sp., 394–395

CChF1, 393

Life cycle assessment (LCA), 414–415, 421–423, 431t

of biogas upgrading technologies, 423–440

software available for, 424t

Life cycle impact assessment (LCIA), 421

Life cycle inventory (LCI), 421

Light intensity, 392–393

Limnothrix planktonica, 484

Linear driving force model (LDF model), 111

Liquefied biogas (LBG), 161

Liquid distribution and redistribution, 86

Liquid electrolyte devices, 229–230

Liquid film flux, 341

Liquid-phase electroreduction, 229–230

Loading point of column, 41

Low-temperature membrane-cryogenic hybrid process, 174–175

M

Macropores, 107

Manhattan project, 117–118

Mass action, 109

Mass transfer

coefficient, 341

of H2, 337–343, 339f

mass transfer/concentration polarization, 367–368

Mathematical modeling of PSA, 112–113

MEDAL company, 118

Media pH, 394

Mediated/direct interspecies electron transfer (MIET/DIET), 16

MeGa-StoRE, See Methane gas storage of renewable energy (MeGa-StoRE)

Membrane

basic membrane processes for gas separation, 117–120, 118f

ors, 35, 39

filtration, 304, 305f

gas absorption principle, 126, 126f

membrane configurations and plant design for upgrading biogas, 140–142

membrane materials and structures, 127–135

carbon molecular sieve membranes, 132–134

inorganic membranes for gas separation, 132

mixed-matrix membranes, 134

polymer structures and influence in permeation, 127–131

results of membrane operations with different materials, 135

membrane-based biogas upgrading, 119

membrane-combinations, 166

polymer membrane materials and gas transport properties, 123t

processes, 448–449

recent developments in membrane-based CO2/CH4 separation, 142–149

separation, 74–75, 161, 277, 473–474

techniques, 277–278

theory of transport in gas separation on membranes, 135–140

Membrane electrode assembly (MEA), 194

Membraneless single-chambered systems, 9

Membraneless systems, 237–239

Membranes evaporators or condensers, 49

Mesopores, 107

Metal organic frameworks (MOFs), 103–104

Metal oxides, adsorption on, 301–302

Methanation, 212, 322–323, 326–327

methanation-based biogas upgrading, 323

methanation-based CH4, 324–326

Methane (CH4), 14–15, 87, 211, 223–224, 321–322, 363, 413–414, 467–468

efficiency, 9

enrichment of biogas, 368–374

electron transfer mechanism, 368–370

microbial communities in biocathode for methane enrichment, 370

state-of-the-art bioelectrochemical biogas upgrading, 370–374, 371t

losses, 44

slip and efficiency, 67–68

Methane gas storage of renewable energy (MeGa-StoRE), 478

Methanobacteriaceae, 370

Methanobacterium spp., 479–480, 483

M. congolense, 344

Methanobrevibacter, 370, 483

M. arboriphilus, 348–349

M. marburgensis, 348–349

Methanoclculus spp., 483

Methanococcus spp., 479–480

Methanoculleus spp., 479–480

Methanogenesis, 9, 10–11, 414

Methanogenic activity, 342–343

Methanogens, 10–11, 327, 369

Methanomicrobium spp., 479–480

Methanosaeta, 368

Methanosarcina spp., 479–480

M. barkeri, 348–349

Methanospirillum spp., 479–480

Methanothermobacter spp., 479–480, 483

M. marburgensis, 333

M. thermautotrophicus, 333

2-Methoxyethanol (2ME), 48

Methyl diethanolamine (MDEA), 31, 31, 33–34

Methyl group (CH3), 129

1-Methyl-1-propyl-piperidinium bis(trifluoromethylsulfonyl) imide ((PMPip) (BTA)), 202–203

(Methylamino) propylamine (MAPA), 42–43

Methyldiethanolamine (MDEA), 417

Methylibium sp., 281

Methylotrophic methanogens, 327

Microalgae-based biogas upgrading and concomitant waste water treatment, 387–388

Microalgal removal of H2S, 311–312

MicrobEnergy, 482

Microbial communities in biocathode for methane enrichment, 370

Microbial conversion of CO2 to CH4 on membrane diff, 148–149

Microbial electrocatalysis, 376–377

Microbial electrochemical separation cell (MESC), 374

Microbial electrochemical systems (MES), 365–366

Microbial electrochemical/electrolysis cells (MEC), 205–206, 224–225, 365– 366, 365f

Microbial electrosynthesis systems (MES), 7, 205–206, 210, 365–366, 365f

Microbial fuel cells (MFCs), 205–206, 224–225, 365–366, 365f

Microbial methanation process, 478–479

Microchannel reactors, 203–205

Microorganisms, 9

Micropores, 107

Microporous hydrophobic membranes, 126

Mist eliminator, 86

Mixed-matrix membranes (MMMs), 122, 134

Molecular diffusion, 110

Molecular sieving, 118, 122, 268–269

Mono-ethanol amine, 74–75

Monoethanolamine (MEA), 29, 31, 31, 64–65, 417, 428–429

Monte Carlo Simulation, 423, 425–426

Multitubular fixed-bed reactors, 198–199

Multiwall carbon nanotube reticulated vitreous carbon (MWCNT-RVC), 369– 370

N

N,N-dimethylformamide (DMF), 239–240

n-dodecane, 276

N-doped CNT, 195

n-hexadecane, 276

n-tetradecane, 276

Nafion, 194

Nannochloropsis, 386

N. gaditana, 395, 484

National Biogas and Manure Management Program (NBMMP), 459–460

Natural gas (NG), 117–118

Net present value (NPV), 450

Nickel(Ni), 230–231

Ni-based catalysts, 201

nickel-based materials, 228

Nitrates (NO3−), 307

Nitrogen (N2), 223–224, 308, 321–322, 363–364, 413–414, 415–416, 468

inhibition, 13

oxidation and reduction reaction, 225

Nitrous oxide (NOX), 14–15

Novel liquid absorbents, 47–48

Novel plate PBR, 391–392

Novel polymer adsorbents, adsorption capacity and regeneration of, 275–276

Number of transfer units (NTU), 82–83

O

Octadecane, 202–203

Office of Gas and Electricity Markets (OFGEM), 457–458

“One through” feed gas flow without preheating, 199

Operation and maintenance cost (O&M cost), 314–316

Operational expenses (OPEX), 343–344

Organic loading rate (OLR), 12–13

Organic physical absorption (OPA), 472–473

Organic waste, 428

Oxazolidin-2-one (OZD), 44

Oxidation–reduction reaction (ORR), 9

Oxygen (O2), 321–322, 337, 363–364, 413–414, 415–416, 467–468, 468

Oxygen evolution reaction (OER), 194–195, 228

Ozone layer depletion, 427

P

P2G, See Power to gas (PtG)

Packed columns, 35, 39

reactors, 35–36, 62–63

Packed-bed design parameters, effect of, 80–83

Packing, 80–81

and gas distributor, 83–85

Pall rings, 35–36

Palladium, 195

Partial product recycling, 199

PEBA, 134

PEBAX composite hollow-fiber membranes, 134

Permeability, 122, 136

Permeance, 121, 121–122

Permselectivity of membrane, 121

Peroxymonosulphate, 258

pH, 11–12, 346–347

Phenomenological approach, 138

Phenyl group (C6H5), 129

Phormidium spp., 484

Photobioreactors (PBRs), 386–387

designs for biogas upgradation, 388–392

Photoinhibition effect, 392–393

Photoperiods, 392–393

Photosynthetic biogas upgrading, 385, 389t, 397t

CO2 and H2S removal through photosynthetic-bacterial associated biogas upgradation, 386–387

impact of different process variables in biogas upgradation, 392–396

future prospects, 396–401

microalgae-based biogas upgrading and concomitant waste water treatment, 387–388

photobioreactor designs for biogas upgradation, 388–392

positive attributes of photosynthetic “microalgae” toward biogas upgradation, 385–386

system, 484–485

Photosynthetic photon flux density (PPFD), 392–393

Photosystem I (PSI), 387

Photosystem II (PSII), 387

Phyllobacterium myrsinacearum, 280

Physical absorption, 74–75, 277

by using organic solvents, 299–300

Physicochemical removal technologies, 298–306

absorption process, 298–300

physical absorption by using organic solvents, 299–300

water scrubbing, 298–299, 299f

adsorption process, 300–304

adsorption on metal oxides, 301–302

adsorption onto activated carbon, 300–301

pressure swing adsorption system, 303–304

membrane separation, 304–306

types, 304–305, 305–306

Physicochemical upgrading technologies, 472–474, 475t

Picochlorum sp., 394

Piperazine (PZ), 31, 32–33, 65–66

Plasticization, 140

Plugging, 86–87

Poiseuille diffusion, 110–112

Poiseuille diffusivity coefficient, 110–111

Policy , 462–463

Polycarbonate (PC), 140

Polydimethylsiloxane (PDMS), 202–203, 265

Polyethylene glycol, 74–75

Polyimide (PI), 118

membranes, 135

Polymer adsorbents (PAs), 258

adsorption capacity and regeneration of novel polymer adsorbents, 275–276


Polymer electrolyte fuel cells (PEFCs), 189–190

Polymer electrolyte membrane (PEM), 474–476

electrolysis, 324

electrolyzers, 193–196, 193f

design, 193–196



water electrolysis, 193–194

Polymer structures and influence in permeation, 127–131

Polysulfone (PS), 117–118

Polytetrafluoroethylene (PTFE), 192

Pore model, 111

Pore volume, 104

Porous crystals, 106–107

Porous membranes for gas separation, 122

PostCap AAS system, 48

Postcombustion flue gas CO2 capture, 47

Potassium carbonate (K2CO3), 62, 301, 416

Potassium hydroxide (KOH), 190–191, 191–192, 428–429

Potassium iodide (KI), 301, 303–304

Potassium titanate (K2TiO3), 192

Power consumption, 100

Power to gas (PtG), 187–188, 188, 245–246, 429, 468

bioelectrochemical methanation, 205–211

challenges and future prospects, 211–212

electrocatalytic methanation, 189–205

primary differences between chemical and biological methanation, 190t

technology for methanation, 474–482

catalytic/thermochemical methanation, 476–478

chemoautotrophic methanation, 478–482

Power to hydrogen (P2H), 429–430

Pressure, 78–79, 102–103

equalization, 101

range, 101

Pressure swing adsorption (PSA), 93–94, 98–99, 298, 363–364, 417–418, 448– 449, 468, 473

adsorbents, 103–107

adsorption isotherm, 107–109

adsorption kinetics, 109–112

advantages, 94

chronological order of historical developments, 94f

mathematical modeling, 112–113

parameters influencing, 99–107

design parameters, 101–103

process performance indicators, 100

sizing, 102

steps, 99f

system, 303–304

Pressurization time, 101

Pretreatment methods, 258–259

Primary amines, 32–33

Process optimization, 49–50

Productivity, 100

Propionibacterium sps., 11–12

Proton exchange membrane (PEM), 189–190

Pseudanabaena sp., 392

Pseudomonas, 280

P. aeruginosa, 280

PTC, See Renewable Electricity Production Tax Credit (PTC)

Pure Knudsen diffusion coefficient, 110

Purge time, 102

Purge-to-feed ratio, 103

R

Random packing, 35–36, 80–81

examples, 36f

technical data, 37t

Raschig rings, 35–36

Reactor

configurations, 330–337

ex situ biomethanation, 332–337

in situ biomethanation, 331–332

design, 193, 199–200

ReCiPe 2016 midpoint (H) method, 426–427

Refrigeration, 259–260

Renewable Electricity Production Tax Credit (PTC), 457

Renewable energy (RE), 187–188

generation targets, 453–454

renewable energy-related policies and regulations, 453–454

Renewable Energy Sources Act of 2000 (RESA), 456

Renewable energy targets (RETs), 453–454

Renewable heat incentive (RHI), 455

Renewable portfolio standard (RPS), 457

Renewables Obligation (RO), 457–458

Research and development (R&D), 31, 479

applications and challenges for, 18–19

Reverse water-gas shift reaction (RWGS), 229

Robeson limit, 140

Robeson plot, 122

Rubbery polymers, 127–128

transport through, 135–137

Ruddlesden–Popper (RP ), 232

S

Sabatier process, 326

Sandwich-type variants, 231–232

Sarcosine, 66

Scenedesmus sp., 386, 392, 393, 395

Scrubbing

column internals, 83–86

demister or entrainment eliminator or mist eliminator, 86

liquid distribution and redistribution, 86

packing and gas distributor, 83–85

factor, 37–39

Secondary amines, 32–33

Selective layer, 123

Selexol, 74–75, 276–277

Sewage treatment plants (STPs), 255–257

Sieve trays, 37

Silanes, 255

Silanols, 255

Silica gel (SG), 259

adsorption capacity and regeneration of, 267–268


Silicone oils, 277

Siloxanes, 255, 363–364, 416–417, 467–468

combined methods for volatile organic silicon compounds removal from biogas, 281

comparison of methods for reducing the content of volatile organic silicon compounds in biogas, 281–285

D4 and D5, 257

methods for reducing content of volatile organic silicon compounds in biogas, 258–281

Skin layer, 123

Slurry bubble column reactors (SBCR), 202

Smooth nonporous pore walls, 111

Sodium hydrosulfide (NaHS), 310–311

Sodium hydroxide (NaOH), 190–191, 191–192, 428–429

Sodium sulfide (Na2S), 310–311

Solid membranes, 305–306

Solid oxide electrolysis cell (SOEC), 189–190, 226–227, 324, 474–476

Solid oxide electrolyzers (SOEs), 196–198, 228

design, 196–197


reactor configurations, 197

Solid polymer electrolysis, 193–194

Solid-oxide devices, 228–229, 237–239

Solid-oxide fuel cell (SOFCs), 226–227

Solubility

of biogas components in water, 77–78

of CO2 in water, 77

Solution-diffusion mechanism, 122

Solvent selection, 32–33

Spirulina, 386, 386–387, 394

S. platensis, 311

Spray columns, 35, 39

Squalane, 202–203

Sterically hindered amines, 32–33

Structured packings, 35–36, 80–81

Subsidies, 452, 453t

Substrate load, 12–13

Succinic acid (SA), 421

Sulfothane process, 310

Sulfur dioxide (SO2), 296

Sulfur oxide (SOX), 14–15

Sulfur-oxidizing nitrate-reducing microorganisms (SO-NR microorganisms), 308

Supersonic separation, 420

Sustainable energy management, 187–188

Swing adsorption technologies, 95–99

efficiency and adsorbents, 96t

Syngas, 228

Synthetic polymer resins, 274–275

Syntrophic acetateoxidizers (SAO), 329

T

Tafel reaction, 190–191

Tax credits, 455

Tax exemptions, 455

Techno-economic studies in anaerobic digestion, 449–452

biogas utilization, 451–452

feed stocks, 450

gas purification technology, 450–451

subsidies, 452

Temperature, 12, 79, 343–344, 394–395

Temperature swing adsorption (TSA), 42–43, 95, 95–97 See also Pressure swing adsorption (PSA)

Thermally rearranged mixed-matrix membranes (TR mixed-matrix membranes), 122

Thermophilic biomethanation systems, 343–344

Thin-film flat-plate PBR, 391–392

Thiobacillus sp., 312, 314, 394

T. denitrificans, 308, 309

T. ferrooxidans 9, 314

T. thioparus, 309

Thiomicrospira denitrificans, 309

THIOPAQ O&G process, 310

THIOPAQ process, 310, 311

Three-phase reactor, 202–203

design, 202


3 phase methanation (3PM), 202

Time cycle, 101–102

Transport equations through glassy polymers, 137–140

Transport fuel, 468–471

Tray columns, 35, 37

Tray perforations, 37

Trickling liquid velocity (TLV), 307

Turbulence, 339

Two-dimensional fixed-bed tube reactor, 200

Two-film theory, 339f, 340

U

Ultra-thin skin CMS (ULT-CMS), 173

Uncertainty analysis, 425–426

Upgrading, 223–224

V

Vacuum pressure swing adsorption (VPSA), 97–98

Vacuum swing adsorption (VSA), 95, 97–98

Valve trays, 37

Viscous diffusion, 110–112

Volatile fatty acids (VFAs), 4, 328–329

Volatile methylsiloxanes (VMSs), 255

adsorption capacity tests

for AAs, 273t

for granulated ACs, 263t

for PAs, 275t

for SGs, 267t

for ZEs, 272t

mutual displacement of volatile methylsiloxanes from activated carbon, 264

physical parameters of granular ACs, 261t, 266t

Volatile organic silicon compounds (VOSCs), 255, 255–257, 282t

comparison of methods for reducing the content of VOSCs in biogas, 281–285

methods for reducing content of volatile organic silicon compounds in biogas, 258–281

adsorption methods, 260–276

pretreatment methods, 258–259

refrigeration and freezing methods, 259–260

nomenclature and basic properties, 256t

removal, 257–258

W

Waste

biorefinery, 17

management policies, 454

Wastewater treatment (WWT), 385

Water (H2O), 74–75, 223–224, 468

effects of operating parameters on CO2 removal in water scrubber, 78–80

electrolysis, 324

flow rate, 79

removal, 416

solubility of biogas components in, 77–78

as solvent for gases, 76–77

water-lean solvents/nonaqueous amine solvents, 48

Water scrubbing, 161, 298–299, 298, 299f, 363–364, 417, 472–473

for biogas upgrading, 57–58

absorption configurations, 62–64

absorption methodologies, 58–62

chemical promoters in water absorption, 64–66

energy consumption, 66–67

methane slip and efficiency, 67–68

challenges and future directions, 86–87

factors affecting biogas upgrading in, 78–83

technology, 75

Weber-Morris model, 111

Wobbe index, 321–322, 384–385

Wood–Ljungdahl pathway, 10

X

XAD polystyrene resins, 274–275

Y

Yttria-stabilized zirconia (YSZ), 197, 230–231

Z

Zeolites (ZEs), 106, 132, 258

adsorption capacity and regeneration of, 269–271


classification and parameters, 270t

physical parameters, 271t

Zero-gap cell, 193

Zinc oxide (ZnO), 301

ZSM-5, 106

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