Emerging Technologies and Biological Systems for Biogas Upgrading 0128228083, 9780128228081

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Emerging Technologies and Biological Systems for Biogas Upgrading

Emerging Technologies and Biological Systems for Biogas Upgrading

Edited by

Nabin Aryal Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark

Lars Ditlev Mørck Ottosen Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark

Michael Vedel Wegener Kofoed Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark

Deepak Pant Separation and Conversion Technology, Flemish Institute for Technological Research (VITO), Mol, Belgium

Academic Press is an imprint of Elsevier 125 London Wall, London EC2Y 5AS, United Kingdom 525 B Street, Suite 1650, San Diego, CA 92101, United States 50 Hampshire Street, 5th Floor, Cambridge, MA 02139, United States The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, United Kingdom Copyright © 2021 Elsevier Inc. All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Details on how to seek permission, further information about the Publisher’s permissions policies and our arrangements with organizations such as the Copyright Clearance Center and the Copyright Licensing Agency, can be found at our website: www.elsevier.com/permissions. This book and the individual contributions contained in it are protected under copyright by the Publisher (other than as may be noted herein). Notices Knowledge and best practice in this field are constantly changing. As new research and experience broaden our understanding, changes in research methods, professional practices, or medical treatment may become necessary. Practitioners and researchers must always rely on their own experience and knowledge in evaluating and using any information, methods, compounds, or experiments described herein. In using such information or methods they should be mindful of their own safety and the safety of others, including parties for whom they have a professional responsibility. To the fullest extent of the law, neither the Publisher nor the authors, contributors, or editors, assume any liability for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions, or ideas contained in the material herein. British Library Cataloguing-in-Publication Data A catalogue record for this book is available from the British Library Library of Congress Cataloging-in-Publication Data A catalog record for this book is available from the Library of Congress ISBN: 978-0-12-822808-1 For Information on all Academic Press publications visit our website at https://www.elsevier.com/books-and-journals Front Cover Picture: The Aarhus University Research Biogas Plant. (Credit Dr. Michael Vedel Wegener Kofoed).

Publisher: Candice Janco Acquisitions Editor: Peter Adamson Editorial Project Manager: Alice Grant Production Project Manager: Nirmala Arumugam Cover Designer: Mark Rogers Typeset by MPS Limited, Chennai, India

Contents List of contributors Foreword Preface

xv xix xxi

Part I Introduction

1

1.

Status of biogas production and biogas upgrading: A global scenario

3

J. Shanthi Sravan, Athmakuri Tharak and S. Venkata Mohan 1.1 1.2 1.3 1.4

Introduction State-of-the-art of biogas production and upgradation Recent trends in biogas utilization: A global prospective Anaerobic digestion 1.4.1 Mechanism of anaerobic digestion 1.4.2 Factors affecting biogas production 1.5 Biohythane 1.6 Electrochemically induced biogas upgradation 1.6.1 Conductive materials in biogas upgradation 1.7 Challenges and way forward Acknowledgments References

Part II Physiochemical upgrading systems 2.

Chemical absorption—amine absorption/stripping technology for biogas upgrading

3 4 6 8 9 11 14 16 17 18 19 19

27 29

Alastair James Ward, Lu Feng and Henrik Bjarne Møller 2.1 Introduction 2.2 Process fundamentals 2.2.1 Amine chemistry 2.2.2 Amine selection 2.2.3 Process description and technology 2.2.4 Energy consumption

29 31 31 32 34 42

v

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3.

Contents

2.2.5 Operational problems and emissions 2.2.6 Economic considerations 2.3 Research and development directions 2.3.1 Novel liquid absorbents 2.3.2 Water-lean solvents/nonaqueous amine solvents 2.3.3 Amine-functionalized solid sorbents 2.3.4 Process optimization 2.4 Conclusions and future perspectives References Further reading

43 45 47 47 48 48 49 50 50 55

Water scrubbing for biogas upgrading: developments and innovations

57

Valerio Paolini, Patrizio Tratzi, Marco Torre, Laura Tomassetti, Marco Segreto and Francesco Petracchini 3.1 Introduction 3.2 Absorption methodologies 3.2.1 Absorption in water (water scrubbing) 3.2.2 Absorption in NaOH solutions (alkaline scrubbing) 3.2.3 Absorption in K2CO3 solutions (hot potassium carbonate) 3.3 Absorption configurations 3.3.1 Packed column reactors 3.3.2 Hollow fiber membrane contactors 3.4 Chemical promoters in water absorption 3.5 Energy consumption 3.6 Methane slip and efficiency 3.7 Conclusions References

4.

Factors affecting CO2 and CH4 separation during biogas upgrading in a water scrubbing process

57 58 58 60 61 62 62 63 64 66 67 68 68

73

Rimika Kapoor, Pooja Ghosh and Virendra Kumar Vijay 4.1 4.2 4.3 4.4 4.5 4.6

Introduction Approaches for CO2 removal from biogas Water scrubbing technology Water as a solvent for gases Solubility of biogas components in water Factors affecting biogas upgrading in water scrubbing process 4.6.1 Effects of operating parameters on CO2 removal in water scrubber 4.6.2 Effect of packed-bed design parameters

73 74 75 76 77 78 78 80

Contents

5.

vii

4.7 Scrubbing column internals 4.7.1 Packing support and gas distributor 4.7.2 Liquid distribution and redistribution 4.7.3 Demister or entrainment eliminator or mist eliminator 4.8 Major challenges and future directions 4.9 Conclusion Acknowledgments References

83 83 86 86 86 87 88 88

Recent developments in pressure swing adsorption for biomethane production

93

Goldy Shah, Shivali Sahota, Virendra Kumar Vijay, Kamal K. Pant and Pooja Ghosh

6.

5.1 Introduction 5.2 Types of swing adsorption technologies 5.2.1 Temperature swing adsorption 5.2.2 Electric swing adsorption 5.2.3 Vacuum swing adsorption 5.2.4 Pressure swing adsorption 5.3 Parameters influencing pressure swing adsorption 5.3.1 Process performance indicators 5.3.2 Design parameters 5.3.3 Adsorbents 5.4 Adsorption isotherm 5.5 Adsorption kinetics 5.5.1 Molecular diffusion 5.5.2 Knudsen diffusivity 5.5.3 Poiseuille diffusion or viscous diffusion 5.6 Mathematical modeling 5.7 Conclusion and future perspectives References

93 95 95 97 97 98 99 100 101 103 107 109 110 110 110 112 113 113

Membrane-based technology for methane separation from biogas

117

Birgir Norddahl, M.C. Roda-Serrat, M. Errico and K.V. Christensen 6.1 Introduction: how the basic membrane processes for gas separation have evolved 6.2 Basic terms of gas separation on membranes 6.3 Membrane materials and structures 6.3.1 Polymer structures and their influence in permeation 6.3.2 Inorganic membranes for gas separation

117 120 127 127 132

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Contents

6.3.3 Carbon molecular sieve membranes 6.3.4 Mixed-matrix membranes 6.3.5 Results of membrane operations with different materials 6.4 Theory of transport in gas separation on membranes 6.4.1 Transport through rubbery polymers 6.4.2 Transport equations through glassy polymers 6.5 Membrane configurations and plant design for upgrading biogas 6.6 Recent developments in membrane-based CO2/CH4 separation 6.6.1 Biogas upgrading by cryogenic and hybrid cryogenic-membrane separation 6.6.2 Biogas upgrading by absorption and hybrid absorption-membrane 6.6.3 Microbial conversion of CO2 to CH4 on a membrane diffuser 6.7 Summary and outlook 6.8 Future developments References

7.

Cryogenic techniques: an innovative approach for biogas upgrading

132 134 135 135 135 137 140 142 143 145 148 150 151 152

159

Francisco Manuel Baena-Moreno, Luz M. Gallego, Fernando Vega and Benito Navarrete 7.1 Introduction 7.2 Cryogenic biogas upgrading 7.2.1 Cryogenic distillation 7.2.2 Cryogenic packed-bed technology 7.3 Cryogenic hybrid systems 7.3.1 Cryogenic-absorption combination process 7.3.2 Cryogenic-adsorption synergized process 7.3.3 Potential combination of cryogenic and membrane processes 7.3.4 Cryogenic-hydrate processes 7.4 Cryogenic-membrane processes 7.5 Full-scale experiences and technoeconomic studies 7.6 Comparison of documented technologies 7.7 Conclusions and future perspectives Appendix I: Conversion factor for unit transformations Appendix II: State forms for CO2 and CH4 as a function of temperature and pressure Acknowledgments References

159 160 162 163 166 167 168 170 170 171 176 177 179 180 180 181 181

Contents

8.

Power-to-gas for methanation

ix 187

Anirudh Bhanu Teja Nelabhotla, Deepak Pant and Carlos Dinamarca

9.

8.1 Introduction 8.2 Electrocatalytic methanation 8.2.1 Alkaline electrolyzers 8.2.2 Polymer electrolyte membrane electrolyzers 8.2.3 Solid oxide electrolyzers 8.2.4 Fixed-bed methanation reactors 8.2.5 Fluidized bed methanation reactors 8.2.6 Three-phase reactor 8.2.7 Micro(channel) reactors 8.3 Bioelectrochemical methanation 8.3.1 Direct electron transfer 8.3.2 Biocathodes 8.3.3 Reactor configurations 8.4 Challenges and future prospects References

187 189 190 193 196 198 200 202 203 205 207 209 210 211 213

Electrochemical approach for biogas upgrading

223

Grzegorz Pasternak 9.1 Introduction 9.2 Faradaic and energy efficiency 9.3 Electroreduction of CO2 9.3.1 Basic considerations 9.3.2 Reactor and process design 9.4 Electrochemical oxidation of H2S 9.4.1 Basic considerations 9.4.2 Reactor and process design 9.5 Biogas upgrading approach and its challenges 9.5.1 CO2 electroreduction 9.5.2 H2S oxidation 9.5.3 Biogas and scale-up approaches 9.6 Concluding remarks and perspectives Acknowledgments References

10. Siloxanes removal from biogas and emerging biological techniques

223 226 226 227 230 236 237 237 241 241 243 244 246 247 247

255

Kazimierz Gaj 10.1 Introduction

255

x

Contents

10.2 Methods for reducing the content of volatile organic silicon compounds in biogas 10.2.1 Pretreatment methods 10.2.2 Refrigeration and freezing methods 10.2.3 Adsorption methods 10.2.4 Absorption methods 10.2.5 Membrane techniques 10.2.6 Biological methods 10.3 Combined methods for volatile organic silicon compounds removal from biogas 10.4 Comparison of the methods for reducing the content of volatile organic silicon compounds in biogas 10.5 Conclusions and future perspective References

258 258 259 260 276 277 278 281 281 286 288

Part III Biological upgrading systems

293

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

295

Anish Ghimire, Raju Gyawali, Piet N.L. Lens and Sunil Prasad Lohani 11.1 Introduction 11.2 Technologies for removal of biogas contaminants 11.3 Physicochemical removal technologies 11.3.1 Absorption process 11.3.2 Adsorption process 11.3.3 Membrane separation 11.4 Ex situ removal using sulfur-oxidizing microorganisms 11.4.1 Biological air filtration 11.4.2 Microalgal removal of H2S 11.5 In situ H2S removal 11.5.1 In situ microaeration 11.5.2 Dosing iron salts/oxides into the digester 11.6 Combined chemical-biological processes 11.7 Comparison of H2S removal techniques 11.8 Conclusions References

12. Biological upgrading of biogas through CO2 conversion to CH4

295 297 298 298 300 304 306 306 311 312 312 313 314 314 316 317

321

Michael Vedel Wegener Kofoed, Mads Borgbjerg Jensen and Lars Ditlev Mørck Ottosen 12.1 Biogas upgrading 12.2 Hydrogen generation and utilization

321 324

Contents

12.3 Methanation 12.4 Microbial basis for biomethanation 12.4.1 Methanogens 12.4.2 Processes in anaerobic digestion 12.5 Reactor configurations 12.5.1 In situ biomethanation 12.5.2 Ex situ biomethanation 12.6 Factors controlling biomethanation 12.6.1 Mass transfer of H2 12.6.2 Temperature 12.6.3 Growth requirements 12.6.4 pH and CO2 12.6.5 Bacterial interaction and competition 12.7 Reactor design for biological methanation 12.7.1 Continuous stirred tank reactor 12.7.2 Trickle-bed reactors 12.8 Future perspectives and applications 12.9 Conclusions Abbreviations list References

13. Bioelectrochemical systems for biogas upgrading and biomethane production

xi 326 327 327 328 330 331 332 337 337 343 344 346 348 350 350 352 354 356 357 357

363

Nabin Aryal, Lars Ditlev Mørck Ottosen, Anders Bentien, Deepak Pant and Michael Vedel Wegener Kofoed 13.1 Background 13.2 Fundamentals of bioelectrochemical biogas upgrading 13.3 Methane enrichment of biogas 13.3.1 Electron transfer mechanism 13.3.2 Microbial communities in biocathode for methane enrichment 13.3.3 State-of-the-art bioelectrochemical biogas upgrading 13.4 Economical insights 13.5 Prospective and challenges 13.6 Conclusion Acknowledgments References

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

363 364 368 368 370 370 375 376 378 378 378

383

Vijay Kumar Garlapati, Swati Sharma and Surajbhan Sevda 14.1 Introduction 14.2 Positive attributes of photosynthetic “microalgae” toward biogas upgradation

383 385

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Contents

14.3 CO2 and H2S removal through photosynthetic-bacterial associated biogas upgradation 14.4 Microalgae-based biogas upgrading and concomitant wastewater treatment 14.5 Photobioreactor designs for biogas upgradation 14.6 Impact of different process variables in biogas upgradation 14.6.1 Light intensity 14.6.2 Media pH 14.6.3 Temperature 14.6.4 Biogas composition 14.6.5 Gas flow rate 14.7 The future prospects 14.8 Conclusion References

386 387 388 392 392 394 394 395 395 396 401 401

Part IV Policy implications for biogas upgrading

411

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

413

Moonmoon Hiloidhari and Shilpi Kumari 15.1 Introduction 15.2 Biomethanation 15.2.1 Cleaning of biogas 15.2.2 Upgrading of biogas into biomethane 15.3 Brief overview of life cycle assessment 15.4 Life cycle assessment of biogas upgrading technologies 15.5 Conclusions Acknowledgment References Further reading

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

413 415 416 417 421 423 441 441 441 445

447

Dhamodharan Kondusamy, Mehak Kaushal, Saumya Ahlawat and Karthik Rajendran 16.1 Introduction 16.2 Role of techno-economic studies in anaerobic digestion 16.2.1 Feedstocks 16.2.2 Gas purification technology 16.2.3 Biogas utilization 16.2.4 Subsidies

448 449 450 450 451 452

Contents

16.3 Successful policies in anaerobic digestion implementation 16.3.1 Policies and regulations 16.3.2 Renewable energy-related policies and regulations 16.3.3 Agriculture policies and regulations 16.3.4 Waste management policies 16.3.5 Incentives 16.3.6 Policy instruments introduced in various countries as a support to AD industry growth 16.4 Decision-support system for biomethane implantation with techno-economic analysis and policies 16.5 Conclusion References

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

xiii 452 452 453 454 454 455 456 461 462 463

467

Ram Chandra Poudel, Dilip Khatiwada, Prakash Aryal and Manju Sapkota 17.1 Introduction 17.2 Biogas composition and feedstock types 17.3 Biogas upgrading for natural gas grid injection and transport fuel 17.4 State-of-the-art of large-scale biogas upgrading technologies 17.4.1 Physicochemical upgrading technologies 17.4.2 Power-to-gas technology for methanation 17.4.3 Bioelectrochemical system (Cambrian Innovation) 17.4.4 Photosynthetic biogas upgrading system 17.5 Conclusion and future perspective References Index

467 468 468 472 472 474 482 484 485 486 493

List of contributors Saumya Ahlawat Department of Biosciences and Bioengineering, Indian Institute of Technology Guwahati, Guwahati, India Nabin Aryal Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark Prakash Aryal Department of Chemical Engineering, Monash University, Clayton, VIC, Australia Francisco Manuel Baena-Moreno Chemical and Environmental Engineering Department, Technical School of Engineering, University of Seville, Sevilla, Spain Anders Bentien Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark K.V. Christensen Department of Green Technology, Faculty of Engineering, University of Southern Denmark, Odense M, Denmark Carlos Dinamarca Department of Process Energy and Environmental Technology, University of South-Eastern Norway, Porsgrunn, Norway M. Errico Department of Green Technology, Faculty of Engineering, University of Southern Denmark, Odense M, Denmark Lu Feng Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark Kazimierz Gaj Department of Environment Protection Engineering, Wroclaw University of Science and Technology, Wrocław, Poland Luz M. Gallego Chemical and Environmental Engineering Department, Technical School of Engineering, University of Seville, Sevilla, Spain Vijay Kumar Garlapati Department of Biotechnology and Bioinformatics, Jaypee University of Information Technology, Waknaghat, India Anish Ghimire Department of Environmental Science and Engineering, Kathmandu University, Dhulikhel, Nepal Pooja Ghosh Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India Raju Gyawali Nepal Electricity Authority, Government of Nepal, Kathmandu, Nepal Moonmoon Hiloidhari IDP in Climate Studies, Indian Institute of Technology Bombay, Mumbai, India

xv

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List of contributors

Mads Borgbjerg Jensen Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark Rimika Kapoor Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India Mehak Kaushal System Biology for Biofuel Group, International Centre for Genetic Engineering and Biotechnology, New Delhi, India Dilip Khatiwada Division of Energy Systems, Department of Energy Technology, KTH Royal Institute of Technology, Stockholm, Sweden Michael Vedel Wegener Kofoed Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark Dhamodharan Kondusamy Department of Civil Engineering, Indian Institute of Technology Guwahati, Guwahati, India; Institute of Soil, Water and Environmental Science, Agricultural Research Organization, Israel Shilpi Kumari Centre for Energy Studies, Indian Institute of Technology Delhi, New Delhi, India Piet N.L. Lens UNESCO—IHE Institute for Water Education, Delft, The Netherlands Sunil Prasad Lohani Department of Mechanical Engineering, Kathmandu University, Dhulikhel, Nepal S. Venkata Mohan Bioengineering and Environmental Sciences Lab, Department of Energy and Environmental Engineering, CSIR-Indian Institute of Chemical Technology (CSIR-IICT), Hyderabad, India; Academy of Scientific and Innovative Research (AcSIR), Ghaziabad, India Henrik Bjarne Møller Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark Benito Navarrete Chemical and Environmental Engineering Department, Technical School of Engineering, University of Seville, Sevilla, Spain Anirudh Bhanu Teja Nelabhotla Department of Process Energy and Environmental Technology, University of South-Eastern Norway, Porsgrunn, Norway Birgir Norddahl Department of Green Technology, Faculty of Engineering, University of Southern Denmark, Odense M, Denmark Lars Ditlev Mørck Ottosen Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark Deepak Pant Separation and Conversion Technology, Flemish Institute for Technological Research (VITO), Mol, Belgium Kamal K. Pant Department of Chemical Engineering, Indian Institute of Technology Delhi, New Delhi, India Valerio Paolini National Research Council of Italy, Institute of Atmospheric Pollution Research, Monterotondo, Italy 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

List of contributors

xvii

Francesco Petracchini National Research Council of Italy, Institute of Atmospheric Pollution Research, Monterotondo, Italy Ram Chandra Poudel Department of Biological Sciences, University of Bergen, Bergen, Norway Karthik Rajendran Department of Environmental Science, SRM University-AP, Mangalagiri, India M.C. Roda-Serrat Department of Green Technology, Faculty of Engineering, University of Southern Denmark, Odense M, Denmark Shivali Sahota Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India Manju Sapkota Institute of Chemistry, Bioscience and Environmental Engineering, Faculty of Science and Technology, University of Stavanger, Stavanger, Norway Marco Segreto National Research Council of Italy, Institute of Atmospheric Pollution Research, Monterotondo, Italy Surajbhan Sevda Department of Biotechnology, National Institute of Technology Warangal, Warangal, India Goldy Shah Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India Swati Sharma Department of Biotechnology and Bioinformatics, Jaypee University of Information Technology, Waknaghat, India J. Shanthi Sravan Bioengineering and Environmental Sciences Lab, Department of Energy and Environmental Engineering, CSIR-Indian Institute of Chemical Technology (CSIR-IICT), Hyderabad, India; Academy of Scientific and Innovative Research (AcSIR), Ghaziabad, India Athmakuri Tharak Bioengineering and Environmental Sciences Lab, Department of Energy and Environmental Engineering, CSIR-Indian Institute of Chemical Technology (CSIR-IICT), Hyderabad, India Laura Tomassetti National Research Council of Italy, Institute of Atmospheric Pollution Research, Monterotondo, Italy Marco Torre National Research Council of Italy, Institute of Atmospheric Pollution Research, Monterotondo, Italy Patrizio Tratzi National Research Council of Italy, Institute of Atmospheric Pollution Research, Monterotondo, Italy Fernando Vega Chemical and Environmental Engineering Department, Technical School of Engineering, University of Seville, Sevilla, Spain Virendra Kumar Vijay Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India Alastair James Ward Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark

Foreword Global energy demand is increasing to fulfill the growing human population needs, with fossil fuels being the most dominating source. One of the most significant environmental problems associated with fossil fuel use is the emission of greenhouse gases (GHGs), leading to global warming and creating problems related to climate change. Increasing the supply of renewable energy sources would replace fossil sources and significantly limit the dominating carbonintensive fossil fuels in the future energy system. Therefore the development and utilization of renewable energy sources such as solar, bioenergy, wind, hydro, and geothermal are essential to mitigate the environmental problems associated with GHG emissions. Bioenergy, where biomass produced via photosynthesis can be converted to biofuel (biogas), heat, and electricity, is the most widely used form of renewable energy. Biogas can be produced from the anaerobic digestion process and can be utilized as a fuel for cooking, industrial processes, and transportation fuels. Nevertheless, biogas contains significant amounts of carbon dioxide and other constituents such as hydrogen sulfide that have to be removed prior to application as a natural gas substitute. By upgrading the biogas and thereby increasing its methane content, the resulting biomethane can replace natural gas obtained from fossil sources. Recently, emerging technologies for biogas upgrading, such as microbial-based and cryogenic-based technologies, have been developed. However, available information about these technologies is limited. This book represents a milestone by providing the technical knowledge and information on emerging biogas upgrading technologies. This book on Emerging Technologies and Biological Systems for Biogas Upgrading provides fundamental knowledge on anaerobic digestion, the global scenario of biogas production, state-of-the-art information on upgrading, and policy implications for promoting the utilization of upgraded biogas. The book deals with physiochemical upgrading systems with great insight into absorption, scrubbing, membrane separation, electrochemical, and cryogenic techniques. The book furthermore presents technologies currently under development, including biological and bioelectrochemical power-to-gas technologies

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employed for biogas upgrading. The editors have put together a host of highly relevant topics and experts in their respective fields to contribute with thoroughly described chapters. What I like about this book is the information about the current state-of-theart, the practical information, and highly qualified consortium of authors, who have contributed with their knowledge in each chapter, which will be highly beneficial for researchers, university students, biogas developers. and practitioners who are entering into the biogas production and biogas-upgrading field. Ashok Pandey Editor-in-Chief, Bioresource Technology, Elsevier

Preface Biogas is a methane-rich gas produced from biological degradation of biomass. The anaerobic digestion (AD) process has been commercially initiated to produce methane (CH4) from organic waste degradation that significantly contributes to global renewable energy production and consumption. Biogas has an important role in the global carbon cycle and has traditionally been used as an alternative renewable energy source. Especially in developing countries, a large part of the rural population relies on decentralized smallscale biogas digestors for meeting their household energy needs and furthermore for utilizing the digestate from such plants as a source of nutrient-rich fertilizer for their soils. Worldwide uncontrolled solid waste production leads to greenhouse gas (GHG) emissions in the form of carbon dioxide (CO2) and CH4, contributing significantly to climate change. Hence, harvesting biogas from the organic waste stream results in an environmentally sustainable source of renewable energy while reducing GHG emissions. According to the World Biogas Association, the annual biogas production in 2018 exceeded 60.8 billion m3 , of which almost 54% was in Europe. The global production is projected to increase further worldwide, illustrating the worth of biogas, especially in a scenario with dramatic reductions in the consumption of fossil fuels. Biogas predominantly consists of 40% 60% CH4, 60% 40% CO2, and traces of hydrogen sulfide (H2S), ammonia (NH3), hydrogen (H2), oxygen (O2), nitrogen (N2), siloxanes, carbon monoxide (CO), hydrocarbons, and volatile organic compounds. Primarily, the CO2 content in biogas lowers its heating value compared to natural gas, and the presence of other constituents may cause corrosion and salt accumulation on the associated appliances such as boilers, burners, and gas engines. Biogas therefore has to be treated and conditioned to improve its gas heating value and downstream applicability. The gas can ultimately reach gas grid quality by upgrading it even further through removal or conversion of CO2. Recently, biogas upgrading has gained intense attention due to national targets for renewable energy production, environmental concerns, and the need to replace fossil fuels with sustainable fuel alternatives. Several technologies are today commercially available and implemented for biogas upgrading at commercial biogas plants, including water scrubbing, amine scrubbing, pressure swing adsorption, and membrane-based technologies. Among the implemented

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Preface

biogas upgrading technologies, water scrubbing is a widely applied technology that accounts for almost 40% of the total upgrading. These scrubbers upgrade the biogas by removing CO2 from biogas and emitting it to the atmosphere. The focus on reducing emissions from the biogas industry and utilizing biogas as a source of CO2 has spawned the development of alternative technologies for biological, bioelectrochemical, and chemical biogas upgrading. These technologies are at different technological readiness levels but all represent promising solutions for reducing the carbon footprint of the biogas industry and at the same time increasing its importance as a supplier of renewable energy and chemicals. As an example, biological methanation today constitutes a promising technology for converting biogas CO2 to CH4 by the use of electricity from renewable sources through a process that combines carbon capture and utilization with energy conversion (Power-to-X). The development and demonstration of new technologies is not only an academic exercise but also includes heavy industrial involvement. Nonetheless, comprehensive access to technical information on biogas upgrading technology remains limited. To overcome such a gap, this book intends to provide complete technical details on biogas upgrading. Each chapter of the book is designed to give a fully comprehensive and most recent state-of-the-art on different technologies currently in use or under development for biogas cleaning and upgrading. In this book, fundamental principles, state-of-the-art, biogas cleaning, and upgrading technologies for CO2, H2S, and siloxane removal or conversion have been elaborated. The book begins by outlining the global scenario of biogas production and upgrading processes, followed by an insight into physicochemical upgrading systems that have been implemented at an industrial scale. The critical process parameter optimization of absorption/stripping technology, and optimization of cryogenic, membrane, and power-to-gas are discussed. The subsequent chapters describe biological cleaning and conversion of H2S, H2-mediated CO2 conversion, bioelectrochemical conversion of CO2 to CH4, and algal-based photosynthetic biogas upgrading. Finally, the implications of policies for biogas upgrading are provided. In a nutshell, the information in this book brings insights into technologies and processes for biogas treatment and upgrading. This book is aimed at a broad audience, mainly researchers, biogas specialists, academics, entrepreneurs, industrialists, policymakers, and others who wish to know the latest developments and future perspectives of biogas upgradation approaches for the enhancement of the existing digestors, and also discusses the bottlenecks of the various technologies that currently limit scale-up and commercialization. The chapters are written by experts in the field from all parts of the world. Consolidation of the most recent state-of-the-art into an independent chapter for each type of physical, chemical, or biotechnological upgrading system is the main aim of this book. A key point of the book is that it also provides guidance on which procedures should be followed under what

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conditions to get the best results in terms of upgrading. It is our sincere hope that this book will contribute to the necessary transition to environmentally benign and sustainable adoption of biogas in general and biogas upgrading approaches in particular. Though we have tried to be objective in our choice of topics to be covered in this book, some not so common themes which may become important in the future may have been missed out, we will try to cover them in the second edition of the book. This book is intended to have three roles and to serve three associated audiences, namely, the students and research community who will benefit from the lucid explanation of the possible applications of biogas upgradation for the betterment of environment, the policymakers who will find it easier to identify the pros and cons of different upgradation systems, and finally, the industries involved, as it will give them a feeling about the current loopholes (technological possibilities and possibilities for optimization) and ways to fix them. Each chapter begins with a fundamental explanation for general readers and ends with in-depth scientific details suitable for expert readers. The text in all the chapters is supported by numerous clear, illustrative, and informative diagrams, flowcharts, and comprehensive tables detailing the scientific advancements, providing an opportunity to understand the process thoroughly and meticulously. Written in an eloquent style, the book comprehensively covers each point to give the reader a holistic picture of biogas treatment technologies and the future perspective of their use. The book may even be adopted as a textbook for university courses that deal with such courses related to both energy and the environment. Despite the best efforts of the authors and editors, along with extensive checks conducted by many experts in the field of biogas and AD, mistakes may have crept in inadvertently. We would appreciate if readers could highlight these and make comments or suggestions to improve and update the book contents for future editions. Nabin Aryal1, Lars Ditlev Mørck Ottosen1, Michael Vedel Wegener Kofoed1 and Deepak Pant2 1

Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark, Separation and Conversion Technology, Flemish Institute for Technological Research (VITO), Mol, Belgium

2

Chapter 1

Status of biogas production and biogas upgrading: A global scenario J. Shanthi Sravan1,2, Athmakuri Tharak1 and S. Venkata Mohan1,2 1

Bioengineering and Environmental Sciences Lab, Department of Energy and Environmental Engineering, CSIR-Indian Institute of Chemical Technology (CSIR-IICT), Hyderabad, India, 2 Academy of Scientific and Innovative Research (AcSIR), Ghaziabad, India

Chapter outline 1.1 Introduction 1.2 State-of-the-art of biogas production and upgradation 1.3 Recent trends in biogas utilization: A global prospective 1.4 Anaerobic digestion 1.4.1 Mechanism of anaerobic digestion

1.1

3 4 6 8 9

1.4.2 Factors affecting biogas production 1.5 Biohythane 1.6 Electrochemically induced biogas upgradation 1.7 Challenges and way forward Acknowledgments References

11 14 16 18 19 19

Introduction

Biogas is considered one of the most sought-after bioenergy resources across the world to overcome the environmental and energy challenges. It has numerous applications in the household, domestic, transportation, gas grid areas, and as a substrate for platform chemicals generation. Conventionally, biogas upgradation (BU) is performed by physico-chemical (absorption, adsorption, membrane seperation, and cryogenic) and biological (in situ and ex situ) processes which are site/case specific (Baena-Moreno et al., 2020; Kapoor et al., 2019; Vrbova and Ciahotny, 2017; Munoz et al., 2015). Limitations need to be addressed for managing of energy and carbon flux, which could essentially benefit the process efficiency (Aryal et al., 2018; Salihu and Alam, 2015). BU technologies need to focus mainly on the reuse of impurities such as CO2, H2S, and other gases generated from individual Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00002-7 © 2021 Elsevier Inc. All rights reserved.

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PART | I Introduction

processes in the presence of electron donors which are essential to produce methane (CH4) with above 90% productivity (Nelabhotla et al., 2019; Sahota et al., 2018; Scholz et al., 2013). BU can potentially enhance the conversion rate/efficiency of organic substrates with further use of impurities such as CO2, H2S, etc. in the presence of electron donors like H2 and volatile fatty acids (VFAs) towards CH4 production and its applications (Sravan et al., 2020; Nelabhotla and Dinamarca, 2018; Venkata Mohan et al., 2016b; Jiang et al., 2013). Electrochemical interference towards CH4 production, called electromethanogenesis (EM), can influence the direct interspecies electron transfer (DIET) with a focus on creating a microbe
electrode synergy that neutralizes the disruptors influence (from side reactions) on enhancing the biogas production (Zhou et al., 2017; Lovley, 2017; Simon, 2015). BU for centralized and decentralized applications needs to provide specific interventions and strategies that could beneficially influence the microbial electrometabolism and its energy dynamics, while focusing on the techno-economics toward increasing the overall CH4 yields at an industrial scale.

1.2

State-of-the-art of biogas production and upgradation

Biogas is an important source of renewable energy that contributes significantly in terms of overall calorific value (Baena-Moreno et al., 2020; Koonaphapdeelert et al., 2020; Curto and Martin, 2019). Biogenic solid and liquid wastes are degraded in anaerobic digestors for CH4 generation along with other gases and heat generation (Gotz et al., 2016; Prasad et al., 2017; Singhal et al., 2017; Subbarao, 2018; Venkata Mohan et al., 2017). The breakdown of organic matter contributes to various biogas constituents, namely, CH4 (50
70%), CO2 (30
50%), and other trace gases(0.1
3%) (BaenaMoreno et al., 2019; Bharathiraja et al., 2018; Kulkarni and Ghanegaonar, 2020; Maurya et al., 2019). Biogas utilization in the gas grid as energy requires a CH4 content of at least 90%, with decreased CO2, H2, H2S, etc., composition (Angelidaki et al., 2018; Kadam and Panwar, 2017). Usually processes like absorption, membrane seperation, scrubbing, and water washing are used to remove the excess CO2 generated (Adnan et al., 2019; Angelidaki et al., 2018). In situ integrated processes that could potentially upgrade the biogas to CH4 within the anaerobic digestion (AD) system could benefit the energy sector (Sarker et al., 2018; Luo and Angelidaki, 2012). Alternative processes to upgrade biogas to CH4 include utilizing the excessive CO2 in the presence of H2 in the AD systems. This upgraded CH4 could be used to meet the requirements of the gas grid and other energy systems (Fig. 1.1). Globally, biogas is being exploited with great interest as a substitute for natural gas. Hence, increasing the calorific value of biogas by removing CO2 and other trace gases for CH4 upgradation is essential (Kapoor et al., 2019; Sun et al., 2015a). BU by fixation of CO2 and H2 in the presence of redox intermediates to

Status of biogas production and biogas upgrading: A global scenario Chapter | 1 Biogas upgradaon

Waste to biomethane and power grid

5

Biomethane and biogas upgradaon

Anaerobic digeson and power grid

Biomethane and power grid

Biogas and power grid

FIGURE 1.1 Scientometric analysis of the current state-of-the-art on biogas upgradation.

CH4 remains undeveloped, but is gaining interest in the context of renewable energy utilization (Alvarez-Gutierrez et al., 2016; Angelidaki et al., 2019). Commercial biogas upgrading systems using conventional processes can perform only the separation of CH4 from CO2 and require additional integration of individual processes to increase the efficiency of CH4 conversion and to also avoid carbon emissions (Baena-Moreno et al., 2020; Xu et al., 2018; Rodero et al., 2018; Vrbova and Ciahotny, 2017; Yuan et al., 2013). The lower density of H2 requires higher storage capacities, while its transportation and direct utilization as a technology is still under development. Hence, the transformation of H2 to CH4 is considered appropriate and could be considered beneficial for its utilization as a natural gas (Fig. 1.2). Methane is advantageous over H2 due to its higher volumetric energy density and its readily available existant infrastructure for its utilization/storage towards application feasibility (Luo et al., 2012). H2 is readily utilized in lab-scale systems for converting CO2 to CH4 with increased conversion efficiency (Luo et al., 2012; Maegaard et al., 2019; Sun et al., 2015b). The increased H2 utilization influences VFA accumulation because of homoacetogenesis leading to higher acidification (Liu et al., 2016). The major limitations to BU are pH regulation and CO2 utilization, where the pH range needs to be maintained between 6.5
8.5 to increase the CH4 production (Bassani et al., 2015; Luo and Angelidaki, 2012). A biological method of BU is considered as a potential alternative for CH4 production using various microbial genera such as hydrogenotrophic methanogens, acetoclastic methanogens, and microalgae (Meier et al., 2017; Muha et al., 2012). In this process, CO2 and H2 are biologically converted to

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PART | I Introduction

Absorpon technology

Cryogenic separaon Absorpon technology

In situ

Water scrubbing

Biogas upgrading

(liquid gas interacon)

Ex situ (removal of CO impuries from biogas)

Hydrogenotrophic methanogenesis

Membrane technology

Biological technique

FIGURE 1.2 Various biogas upgradation techniques/processes.

CH4 involving the action of hydrogenotrophic methanogens without any additional energy inputs [Eq. (1.1)]. 4H2 1 CO2 -CH4 1 2H2 O

ΔGO 5 2 130:7 kJ=mol:

ð1:1Þ

H2 injection with a stoichiometric ratio of 4:1 between H2 and CO2 during hydrogenotrophic methanogenesis increases CO2 utilization thus leading to a pH increase, which is one of the main influencing parameters for efficient performance of the methanogenic population (Liu et al., 2016; Luo and Angelidaki, 2012; Maegaard et al., 2019; Siegert et al., 2015). Higher alkaline pH values usually limit the methanogenic activity, while CO2 utilization helps overcome the substrate inhibition for autotrophic hydrogenotrophic methanogens towards CH4 production. The methanogenic population in AD systems usually consists of acetoclastic and hydrogenotrophic methanogens, majorly contributing to the CH4 production (Liu et al., 2018; Christy et al., 2014; Luo et al., 2012; Sarkar and Venkata Mohan, 2020). A higher H2 presence helps towards the enrichment of hydrogenotrophic methanogens like Methanomicrobium, Methanoculleus, and Methanobacterium, which relatively increases the rate of methanogenesis (Luo and Angelidaki, 2012; Bassani et al., 2015).

1.3

Recent trends in biogas utilization: A global prospective

Biogas is one of the most important sources of renewable energy produced from anaerobic digestors and could contribute significantly in terms of energy value (Curto and Martin, 2019). Biological CH4 production is produced by dark fermentation, where the organic substrate is converted into

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biogas, biofuels, and other value-added intermediate products. Electrodes and a polarizing microenvironment with applied potential could be beneficial in regulating the microbial metabolism and increasing the substrate conversion rate towards BU (Castellano-Hinojosa et al., 2018; Dou et al., 2018; Liu et al., 2016; Nikhil et al., 2015; Schroder et al., 2015; Zhao et al., 2016). Apart from these, redox intermediates (activated carbon/biochar/magnetic field) act as redox shuttles for electron acceptance, influencing the microbial electrogenic activity towards decreased losses and increasing CH4 recovery. Integration of a polarized microenvironment with dark fermentation, called electromethanogenesis (EM) could be innovative in increasing the CH4 conversion rate, productivity, and calorific value. EM regulates the electron flux with the endogenous or applied potential establishing synergistic redox microbe
electrode and microbe
microbe interactions (Sravan et al., 2020; Modestra et al., 2015a,b). Microbial activity with deprived electrons needs an increased energy conversion rate and therefore could utilize the applied potential to regulate their metabolism, resulting in higher CH4 production (Ren et al., 2019; Villano et al., 2017; Jin et al., 2017). BU/EM systems emphasize electrode material placement, biocatalyst, system design, and operation. Dark fermentation is mainly focused on short-chain carboxylic acids (C2
C6) and alcohol production, with respect to methanogenic microbial suppression losing a highly significant amount of energy in the form of biogas. However, EM stresses streamlining towards higher CH4 production to meet the increasing demands with regulated microbial metabolism, microenvironment, and energetics. Biogas production has recently focused specifically on CO2 capture. Electrochemical CO2 reduction to CH4 selectively orients for BU using bioelectrochemical systems (BES), and is described as power to gas technology (Collet et al., 2017; Stangeland et al., 2017; Zhao et al., 2016; Xu et al., 2014). Overcoming the limitations of DIET and cathode development shows a marked effect on CO2 reduction and CH4 production, involving electroactive microbial catalysis with polarized electrodes and applied potential (Dykstra and Pavlostathis, 2017; Fu et al., 2015; Sravan et al., 2020). An optimum pH of 7.0
7.5 helps to increase the biogas productivity and also favors enrichment of electroactive methanogens. Several studies have approached in the direction of integrating BES with AD for biogas upgradation with increased purity. A polarized microbial environment with the interference of electrodes in the microenvironment favors the conversion of raw biogas components with the support of intermediate acceptors to CH4 (Jiang et al., 2019; Sravan et al., 2020; Venkata Mohan et al., 2014; Xu et al., 2014). In situ developed or applied potential regulates the charge transfer kinetics between the electroactive microbial populations and inert electrode surfaces, enhancing the CH4 production (Ren et al., 2019; Paiano et al., 2019). Applied potential drives the methanogenic metabolic

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PART | I Introduction

Motor

Gas outlet Raw biogas

Biogenic waste (substrate) inlet

Effluent oulow Acid/base inflow

Biogas

CO

CO reducon to methane

Membrane separaon of CO

CH

Baffles

H ,H S

Impeller Substrate

Biomethane (methane 90%–95%)

Anaerobic bacteria Gas grid Raw biogas

CH 50%–75% CO 20%–40% H , H S 5%–10%

Phase I : Biogas producon

Household supply

Industrial supply

Phase II : Biogas upgradaon

FIGURE 1.3 Integration of anaerobic digestors with biogas upgradation to increase methane production.

pathways in order to increase the substrate conversion rates along with productivity (Sarkar and Venkata Mohan, 2020; Sravan et al., 2020; Meier et al., 2017; Liu et al., 2016). Hydrogenotrophic methanogens were studied to increase the CH4 production from endogenous H2 produced by reutilizing it as an additional electron donor, whereas acetoclastic methanogens were involved in the inhibition of H2 towards CH4 production. Different types of feedstocks like municipal solid waste, spent wash, domestic food waste, and C1 gases like CO2 and CO were used in BES to achieve higher BU rates (Jiang et al., 2013; Zhen et al., 2017) (Fig. 1.3).

1.4

Anaerobic digestion

AD is a conventional process for the biological conversion of organic substrate to biogas, mainly CH4 and CO2, with other trace gases along with other value-added products. AD requires a longer operation time to achieve effective substrate removal and value-addition yields. The lower conversion efficiencies during AD (upto 60%) are considered as a disadvantage for the process performance. Recently, BU has gained significance for the improvement of CH4 content in total biogas yields, along with value-addition with AD
BES integration. The biological conversion efficiency of organic substrate in the presence of H2 and CO2 to CH4 occurs at the anode/cathode in BES, and is also called the EM process (Siegert et al., 2015).

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Single-chambered AD and BES have shown improved CH4 efficiency with the presence/enrichment of hydrogenotrophic methanogens that are significantly involved in H2 and CO2 conversion to CH4 (Sarkar and Venkata Mohan, 2020; Sravan et al., 2020). CH4 production was also improved in a single-chambered BES, involving H2 gas recycling at the anode towards higher waste utilization and process intensification. Membraneless single-chambered systems are robust for CH4 production rather than being confined to hydrogen at a slightly acidified pH. AD
BES integration showed a significant increase in CH4 composition in the biogas (Sravan et al., 2020; Liu et al., 2018; Chen et al., 2016).

1.4.1

CO2 1 4H2 -CH4 1 2H2 O

ð1:2Þ

1 2CH3 COO2 1 6H1 1 4e2 -C4 H7 O2 2 1 H 1 2H2 O

ð1:3Þ

Mechanism of anaerobic digestion

The organic substrate during AD is catabolized through anaerobic fermentation and anaerobic respiration under dark conditions in the absence of oxygen as an electron acceptor. Microorganisms conserve energy through an internally balanced oxidation
reduction reaction (ORR) in anaerobic fermentation, whereas in anaerobic respiration it uses nitrates, sulfates, fumarates, etc., as the electron acceptor rather than oxygen. AD involves four steps, i.e., hydrolysis, acidogenesis, acetogenesis, and methanogenesis. The multi-molecular organic substrates are converted into simple, chemically stabilized molecules using H2 and acid as intermediate metabolites, and finally to CH4 and CO2 (Sarkar and Venkata Mohan, 2020).

1.4.1.1 Hydrolysis and acidogenesis During hydrolysis, the microorganisms initially hydrolyze the complex organic polymers to monomers and further ferment them to a mixture of organic acids and alcohols, mainly through the Embden
Meyerhof
Parnas or Entner
Doudoroff pathways (Angelidaki et al., 2018). The hydrolysis rate mainly depends on the particle size, pH, gas diffusion/production, and enzyme adsorption on the waste particles in AD (Appels et al., 2008). Extracellular enzymes belonging to the hydrolases group such as amylase, protease, and lipase produced by specific hydrolytic bacteria are used during hydrolysis. Biochemical pathways end with pyruvate as a key intermediate which is utilized as an internal electron acceptor for the NADH reoxidation for production of VFAs (C2
C6) such as acetate, propionate, butyrate, lactate, valerate, and caproate, along with H2 and formate. In acidogenesis, the acidifying bacteria involving hydrogenation and dehydrogenation convert organic substrates and hydrolysis products to organic acids, alcohols,

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PART | I Introduction

aldehydes, CO2, and H2. Pyruvate during anaerobic respiration can be further oxidized to acetate using acetogenic bacteria for comparatively higher production of oxidized end products (acetate and CO2), which inherently increases the overall ATP yield.

1.4.1.2 Acetogenesis Acetogenesis is performed by a Wood
Ljungdahl pathway involving a phylogenetically diverse bacterial group (acetogens) which is specifically characterized by CO2 reduction to the acetyl-co-enzyme A. The acetylCoA pathway serves towards electron acceptance and carbon assimilation with energy conservation, where the carbon sources are used effectively as electron donors/acceptors for enrichment of autotrophic/heterotrophic bacteria. In acetogens, the electron donors are CO, H2, formate, methyl chloride, pyruvate, lactate, oxalate, etc., and the electron acceptors are CO2, fumarate, nitrate, thiosulfate, dimethyl sulfoxide, pyruvate acetaldehyde, and H1. Acetogenesis directly illustrates the biogas production efficiency in the system, with approximately 70% of CH4 coming from acetate reduction, in which approximately 25% of acetates and 11% of H2 are produced. Hence, it is crucial to target the acetogenic phase for higher CH4 output with an integrated approach towards increasing its energy and calorific value. 1.4.1.3 Methanogenesis Methanogenesis is the final process for CH4 production by methanogenic microorganisms, which convert acidogenesis/acetogenesis products (VFAs, H2, CO2, H2S, and alcohols) into CH4. Methanogens are classified into two groups based on the substrate availability. Hydrogenotrophic methanogens use H2/formate as an energy source in the presence of certain electron donor like alcohols and CO2 to reduce to CH4. Methanogens are either obligate that use only H2/formate as the sole electron donor or more flexible and use other forms of energy sources. Methylotrophic methanogens are a more versatile group that include organic substrates such as VFA, H2, CO2, CO, and alcohols for methanogenesis using the methyl-S-CoM pathway. Hydrogen and interspecies formate transfer happens during methanogenic syntropy where acetate acts as an efficient electron carrier between the syntrophic partners. The electron transfer with direct contact of microbes without production of H2 occurs in the presence of conductive pili and is called DIET (Sravan et al., 2020, Lovley, 2017; Raghavulu et al., 2012). H2 utilization provides optimized conditions for the enrichment of acidogenic bacteria, which produce short-chain carboxylic acids (VFAs) in the acidification phase followed by low H2 production in the acetogenic phase. These conversions produce

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CO2-rich gas which can be effectively converted to CH4. Microbial communities show significant variations in diversity in a short period of time by varying different reactions in single process. Microbial diversity is directly influenced by the operating parameters, namely, substrate and its composition, pH, temperature, pretreatment, and retention time (Sunny and Joseph, 2018; Venkata Mohan et al., 2008).

1.4.2

Factors affecting biogas production

1.4.2.1 Hydrolysis Hydrolysis is a key rate-limiting factor during AD that influences the conversion efficiency of the waste into CH4 and associated products. Efficient pretreatment strategies (physical and chemical) have been mostly focused on overcoming sludge hydrolysis (Venkata Mohan et al., 2008). The various pretreatment strategies directly influence the specificity of products generated by influencing the granular size of the sludge that leads to the enrichment/inactivation of biocatalyst (Sarkar and Venkata Mohan, 2016). Hence, the microbes involve a higher rate of degradation by recycling the carbon and nutrients in the digesters. Anaerobic digesters when integrated with BES influence on increasing the fermentation efficiency by cascading the individual processes towards higher utilization of redox equivalents. The increase in the electrocatalytic rates and electrometabolism shows a significant impact on the microbe
electrode interaction and biofilm formation. 1.4.2.2 pH The optimal pH range is a critical parameter that influences the AD process, thereby affecting the CH4 output and product synthesis. pH represents the hydrogen ion concentration in the digestion medium and its variations have a direct influence on the growth rate and metabolism of the microbial community. A near neutral pH (6.8
7.4) is considered as the ideal pH for the enrichment, growth, and relative abundance of methanogenic microbial community towards increasing the CH4 production (Liu et al., 2018; Sravan et al., 2020). Studies on acidogenesis have confirmed the variations in relative abundance of a particular species at certain pH ranges. pH 6 was found to be more suitable for the growth of Clostridium butyricum and pH 8 for Propionibacterium sp. to dominate and perpetuate their communities during AD. Methanogenesis is another important process during AD that is regulated at an optimal pH range of 6.5
8.2, but in most cases pH 7 is considered as ideal for its production (Liu et al., 2018; Cavinato et al., 2013). The development of these specific microbial communities has greatly influenced the VFA composition, where at the optimal pH range the enzymatic activity of the microorganisms is higher,

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PART | I Introduction

leading to the production of higher amounts of fatty acids (Hajji et al., 2016). The production of VFA in relation to the optimal pH also needs to consider the COD conversion efficiency of the microorganisms which reflects their metabolic activity. A positive correlation between the pH and hydrolysis was also established and is known to influence the conversion efficiency of the organic substrates to products (Venkata Mohan et al., 2016a). Integration of AD-BES processes is by an electrocatalysis mechanism for supplementation of additional electrons to the existing process and decreasing the energy requirements for BU and the synthesis of other products. Therefore, pH needs to be considered as the most influential factor that controls the rate-limiting factors involved during the AD process for higher product synthesis.

1.4.2.3 Temperature Ambient temperature is a requirement of the AD process, where it exhibits faster reaction rates with stability at higher organic loading of substrate for increased biogas production. Variations in temperature result in variations in the profile of the microorganisms, giving them stability to tolerate adverse conditions (Sunny and Joseph, 2018). Thermophilic (55
70 C) and mesophilic (37 C) conditions are the most prevalent conditions involved during AD (Liu et al., 2018; Dobre et al., 2014). Thermophilic conditions have the advantage of higher biodegradability with the provision of heat energy resulting in higher product synthesis. It is more suitable for acidification while inhibiting biogas production. It can influence the effluents from the process which are environmentally susceptible, while lower biogas output and energy inputs affect the overall economics of the process. Mesophilic conditions provide comparatively better stability and microbial abundance, with the capacity to produce higher CH4 yields. Hence, thermophilic conditions are mostly suitable for acidogenesis, and mesophilic conditions for the methanogenesis process. AD microorganisms are directly influenced by optimal temperature changes that can vary the biogas production and product synthesis quantities. The integration of electrochemical processes with conventional fermentation could increase the rate of product synthesis. 1.4.2.4 Substrate load A higher substrate load is a parameter that leads to bacterial inhibition decreases the productivity of the AD process (Jiang et al., 2013; Babaee and Shayegan, 2011). A higher organic loading rate (OLR) increases the rate of hydrolysis/acidification compared with methanogenesis, which eventually leads to increased VFA production and bacterial inhibition (Pasupuleti et al., 2014). The increased VFAs lower the pH, making the microenvironment acidic and thereby negatively influencing the methanogenic microorganisms

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which cannot further convert the VFAs to CH4. Hence effluent recirculation and integration of AD-BES have great potential to decrease the overloading inhibition. Microbial community profiling varies with organic load with Firmicutes being predominant at lower OLR and Gammaproteobacteria, Actinobacteria, Bacteroidetes, and Deferribacteres being observed at higher substrate loads.

1.4.2.5 C/N ratio An optimal substrate load also needs to reflect specific levels of nutrients in the form of carbon to nitrogen (C/N) ratio. Provision of sufficient amounts of nutrients to the microbial community helps in the maintenance of biomass and faster utilization of nitrogen, resulting in higher biogas production. Lower C/N ratios decrease nitrogen inhibition, which is toxic to methanogens and leads to reduced utilization of carbon sources. The presence of nitrogen in the organic substrate benefits as an important element for the synthesis of amino acids and proteins, while proteins are further converted to ammonia, which helps in maintaining a favorable pH microenvironment for microorganisms. A higher nitrogen content causes toxic effects, while lower quantities of nitrogen cause nutrient limitation (Khalid et al., 2011). The C/N ratio range of 20:1 to 35:1 is considered optimum, and the ratio of 25:1 is considered ideal for the AD process (Christy et al., 2014; Ellabban et al., 2014). The C/N ratio of 25:1 resulted in three-fold higher biogas production when compared to a C/N ratio of 15:1. Hence, an ideal substrate load with a specific C/N ratio would help in higher biogas production by influencing the metabolism of microorganisms involved in the AD process. 1.4.2.6 Hydraulic retention time The hydraulic retention time (HRT) critically influences an increase in biogas production/upgradation. It indicates the period of time at which the productivity could start to decline, while the organic fermentable substrate remains in the anaerobic digester. Increased HRT will require a large digester volume, increasing the overall operational cost, while a shorter HRT will remove the active bacterial population (Sreekrishnan et al., 2004). Maximum CH4 production and its upgradation essentially occur at optimized HRTs. The optimized HRT mainly depends on the type of biocatalyst (mixed or pure culture) and the OLR. A shorter retention period leads to VFA accumulation that causes severe fouling, resulting in decreased biogas production, whereas if the retention time is longer, the biogas components are not utilized effectively, resulting in decreased biogas production (Chen et al., 2016; Dobre et al., 2014). HRT also depends on the reactor size and volume (L/D ratio), where in lab-scale operations, the HRT is much less because of the small reactor size, but in contrast the

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PART | I Introduction

HRT in centralized biogas systems is high to due pilot-level operations. Hence, HRT needs to be considered for BU for increased CH4 production during an integrated process.

1.5

Biohythane

Biohydrogen (H2) is a clean and sustainable energy-dense fuel which is biologically produced during anaerobic fermentation (AF)/AD, photofermentation, biophotolysis, and integration of these individual processes (Pasupuleti et al., 2014; Venkata Mohan et al., 2009). The global economy is expected to rely on H2 as a primary source of energy with zero carbon emissions and high energy-carrying capacity (Venkata Mohan and Sarkar, 2017; Sharma and Ghoshal, 2015). H2 is produced by obligatory acetogenic bacteria using renewable organic sources (Roy and Das, 2016; Sarkar and Venkata Mohan, 2020; Venkata Mohan et al., 2009). Biogenic waste with organic fraction act as a carbon and energy source for the microorganisms for H2 production. The production of H2 by the acidogenic/dark-fermentation process is at a higher rate and is a versatile process which is light independent, and converts biogenic organic wastes predominantly to VFAs (acetic, propionic, and butyric acids) along with simultaneous H2 production (Sarkar and Venkata Mohan, 2017; Dahiya et al., 2018). H2 production through dark fermentation as an individual process has certain limitations. The gaseous energy recovery in terms of only H2 is not sufficient for its commercial viability and application, where only 20
30% of total gaseous energy is recovered through H2 production (Sarkar and Venkata Mohan, 2016; Edison, 2014; Bauer et al., 2013). Integrated processes need to be commercialized for the economic feasibility of H2 production via dark fermentation which is worthy of commercialization, where it could be essentially integrated with AD. AD processes are easy to scale up, and the integration of these two processes can lead to .50
60% gaseous energy recovery (Sen et al., 2016). The integration of AD and dark fermentation processes would also help to decrease the operational cost. The development of such processes would lead to decentralized use of both H2 and CH4. Hence, integrated processes with varied microorganisms and individual capabilities need to be exploited to overcome the disadvantages of individual processes and to enhance the system energetics. Biohythane, an alternate renewable biofuel, can be potentially produced when H2 and CH4 are mixed in appropriate ratios with a blend of 75 6 90% (v/v) CH4 and 10 6 25% (v/v) H2 to make an alternative to fossil-based fuels (Pasupuleti and Venkata Mohan, 2015; Sarkar and Venkata Mohan, 2016). H2 and CH4 are the most widely used biofuels due to their high calorific values of 143 and 55 kJ/g, respectively (Pasupuleti et al., 2014; Sharma and Ghoshal, 2015; Edison, 2014). Biohythane is H2-enriched CH4 that has the scope to be a good alternative

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to the increasing demands for compressed natural gas (CNG) as an engine fuel. H2 is considered as a clean energy fuel since it does not release even a small fraction of CO2 into the atmosphere during combustion (Roy and Das, 2016). On the other hand, CH4 combustion generates greenhouse gases such as CO2. Also, utilization and combustion of both CH4 and H2 do not show any evidence for the release of NOX (nitrous oxide) or SOX (sulfur oxide). The lower ignition power of CH4 and highly flammable nature of H2 are usually considered as drawbacks while individually using them as vehicle fuels. The individual limitations of H 2 and CH4 can be overcome with this blending in optimized proportions to form biohythane (Sarkar and Venkata Mohan, 2016; Dahiya et al., 2018). Its appropriate blending makes it a fuel that is clean with a good calorific efficiency (Pasupuleti and Venkata Mohan, 2015; Sarkar and Venkata Mohan, 2017; Sen et al., 2016). Biohythane has numerous practical applications as a vehicular fuel and is comparatively advantageous over CNG. Its high H2-reducing power increases the combustion rate and burning capacity of CH4 (Roy and Das, 2016; Moreno et al., 2012). It is an eco-friendly fuel due to its advantage of reducing the impact of greenhouse gas emissions on the environment, while the H2 presence helps to decrease the carbon in the gas mixture. Biohythane production was evaluated from the lab to semi-pilot scale, while more recently the blending of flammable H2 gas with CH4 grid injections as a technology was established on a large scale for biohythane production in vehicle fuel plants. Biohythane mimics the hydrogen-enriched compressed natural gas (H-CNG) composition. H-CNG supplemented with emission-free hydrogen (H2) has application feasibility in both residential (heating and cooking) and transport sectors (vehicle engines) as fuel with lower emissions being the prime advantage (Talibi et al., 2017). H-CNG helps in increasing the flammability limits, speed propagation, pressure rise, and deflagration index when compared to CNG. European Union project “NATURALHY” studied the blending of H2 to natural gas for clean combustion of CNG with efficient calorific value and lower ignition energy requirement (Cinti et al., 2019). H-CNG in recent times is receiving significant prominence as an energy carrier, due to its application flexibility with the existing engines (Khab et al., 2019; Miao et al., 2011). This application flexibility could help in injecting H2 in the existing gas pipelines/natural gas grids for both industrial and household purposes related to the transition toward hydrogen economy and would be a beneficial factor for commercial application of the technology at the root level. The molecular H2 as a blend (25–50%) in CNG is currently being derived from fossils in the market which could be environment-impacting with significant emissions being produced (Cetinkaya et al., 2012). Hence, the alternate biological (AD)/bioelectrochemical (BES) and their integrated processes producing green/ low-carbon H2 need to be considered to further suit the advantages of H-CNG. Regulation of process parameters such as pretreatment, pH, microenvironment,

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PART | I Introduction

and specific bacterial enrichment needs to be considered to substantially produce higher H2 than CH4.

1.6

Electrochemically induced biogas upgradation

BU involves a synergy of microbial interactions that show a regulatory influence on electron flux, resulting in the conversion/utilization/reduction of CO2 for CH4 production (Deng et al., 2020; Sarkar and Venkata Mohan, 2020; Sravan et al., 2020). Fermentation redox intermediates (H2, CO2, VFAs, etc.) from the microbial metabolic side reactions counter the targeted end-product due to endogenous losses. BU depends on the syntrophic interactions between fermentative and methanogenic microorganisms to increase electron transfer via mediated/direct interspecies electron transfer (MIET/DIET) to increase the H2 utilization and other electron carriers and redox intermediates towards enhanced CH4 production (Sravan et al., 2020; Deng et al., 2020; Yang et al., 2019). Microbial interactions for increased electrogenic activity could be triggered for increased performance during AD, with the polarized potential developed due to electrode placement or by the external supplementation of potential towards higher CH4 production, described as electromethanogenesis (EM). Electrode placement or applied potential to a microenvironment influences on increasing the reaction/electron transfer rates with respect to conventional fermentations towards increasing the CH4 content in total biogas (Sravan et al., 2020; Deng et al., 2020; Meier et al., 2017). It influences an increase in the microbial electrocatalysis while controlling electron flux, energy utilization, and system buffering for CO2 conversion in the presence of H2 and VFA to CH4. The EM strategy in the presence of electrodes or applied potential helps in efficiently neutralizing/reducing the overpotentials and electrochemical losses to overcome the limitations of BU. Hydrogenotrophic methanogenesis aids in the in situ H2 utilization or reducing equivalents (e2 and H1) for CO2 reduction to enhance additional CH4 production (Sravan et al., 2018, 2020). Hydrogenotrophic methanogens directly utilize H1 and e2 with the use of lower activation energy as an electrocatalytic activity for CO2 reduction to form CH4. The use of lower activation energy aids in establishing efficient microbe
electrode interactions to increase the CO2 reduction. Homoacetogens directly involve CO2 and H2 reduction to acetic acid, which is utilized further for CH4 production (Villano et al., 2010). Syntrophic interactions of homoacetogenic and hydrogenotrophic bacteria towards CH4 utilize H2 as an electron donor, while inhibiting medium/long-chain fatty acid formation with regulated microbial metabolism (Rader and Logan, 2010). 4H2 1 2CO2 -CH3 COOH 1 2H2 O

ð1:4Þ

CH3 COOH 1 2H2 O-CH4 1 CO2

ð1:5Þ

Status of biogas production and biogas upgrading: A global scenario Chapter | 1

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EM also depends on the metabolic microenvironment which is vital to understand the metabolic pathways for the targeted products (Sarkar and Venkata Mohan, 2020; Jiang et al., 2019; Paiano et al., 2019). The anodic metabolic function of the BES effectively contributes to energy generation with respect to substrate oxidation (Sravan et al., 2020; Moscoviz et al., 2016; Nealson and Rowe, 2016). Enriched microbes under the specified microenvironment increase the process efficiency by creating equilibrium between substrate oxidation and oxygen utilization (Nealson and Rowe, 2016; Raghavulu et al., 2012). EM in synergy with microbe
electrode interactions and the specific microenvironment helps in regulating metabolite biosynthesis for CH4 production and could be considered as an essential unit operation in the waste biorefinery.

1.6.1

Conductive materials in biogas upgradation

Direct interspecies electrons transfer (DIET), a syntrophic metabolism where free electrons flow from one cell to another through shared physical (microbe-microbe/microbe-electrode) and electrical connections (via conductive pili) without the requirement of reduced electron carriers (redox mediators) like molecular hydrogen or formate. Biogas production from the conventional AD has several rate-limiting factors such as a) accumulation of intermediate compounds like VFA that affects the process forward b) development H2 partial pressure in the digester leads to inhibition of specific co-enzymes of methanogenic bacteria c) inhibition of ammonia traces developed d) washout of the methanogenic biofilm during feeding (Baek et al., 2018; Fagbohungbe et al., 2015). Conducting materials act as efficient redox mediators by shuttling electrons between the syntrophic microorganisms. Conductive materials provide the regulatory influence of electron transfer between the microbe-microbe and microbe-electrode in a specified redox environment in AF/AD for BU. Parent inoculum, pretreatment, pH, microenvironment, etc., critically influence DIET with conductive materials for BU in AD, BES, and integrated systems (Baek et al., 2018; Zhao et al., 2017). The conductive materials mediated DIET with varied concentrations have shown to be highly efficient in the enhancement of CH4 yield. In AD, the transfer of electrons between two different syntrophic microbial communities such as bacteria and archaea is a vital process for methanogens to get control of energy barriers and catabolized complex organics. Amendment of the several conductive materials to AD, accelerate DIET between microbes leading to increased process efficiencies (Cheng et al., 2020). Conductive materials viz., granular activated carbon (GAC), carbon cloth, biochar, reduced graphene, iron conducting materials like magnetite (Fe3O4), carbon nanotubes are being successfully implemented as an approach to improve the AD process for improved CH4 production (Tan et al., 2021; Zhao et al., 2017; Cheng et al., 2020). The higher

18

PART | I Introduction

surface area of the conductive materials favours the growth of the methanogens with tendency to form dense aggregates of biofilm. Higher conductivity and biocompatibility of conductive materials positively influences the DIET between the microbes leading towards process intensification resulting in increased CH4 content.

1.7

Challenges and way forward

BU has the scope and possibility for practical applications with an integrated electrochemical strategy. BU increases the CH4 yields by enhancing the reduction of CO2 during the process (Sravan et al., 2020; Chen and Liu, 2017; Zhao et al., 2016; Beese-Vasbender et al., 2015). The calorific value of biogas is directly proportional to its CH4 content. Hence, increasing the CH4 composition in the biogas increases the energy and economic value, while also decreasing the transportation and storage costs. The increase in CH4 composition in biogas also significantly improves the feasibility and compatibility of its utilization in natural gas distribution (Mamun et al., 2016). CH4 can be used effectively as a vehicle fuel or directly injected into the gas grid for storage due to its high calorific energy content (Persson et al., 2006). The biogas can also be used directly for diverse applications such as for heating using gas-based boilers and cooking using gas stoves and ovens (Harasimowicz et al., 2007). Household supply

Industrial supply

Household

Industries

Biomethane distribuon

Local biogas head unit

Biomethane distribuon

Biogenic waste

Industrial biogas supply head

Anaerobic digeson unit Gas transmission line

Gas transmission line

Biogas upgrading unit

Biogas grid injuncon

Household biogas supply line Industrial biogas supply line

FIGURE 1.4 Biogas grid supply schematic for industrial and household purposes.

Status of biogas production and biogas upgrading: A global scenario Chapter | 1

19

BU applications are also associated with the removal of organic substrates and harmful trace components as an additional benefit during the process. BU with integrated bioelectrochemical strategy for the conversion of CO2 to CH4 has potential feasibility for application in replacing the stripping tower in a water scrubbing unit (Vijayanand and Singaravelu, 2017; Cheng et al., 2009). The integration of electrochemical energy with AD possibly regulates the rate of CO2 utilization with specificity towards product synthesis by nongenetically regulating the microbial metabolism. It also influences the direct electron transfer with effective microbe
electrode interactions during the operation. The syntrophic interactions between the microbes and electrode, through extracellular electron transfer, catalyze the anodic and/or cathodic reactions. The electrochemically driven BU provides a solution for excessive CO2 produced in the process to be directly transformed into CH4 rather than its separation from the biogas, significantly increasing the productivity and energy value of biogas plants. The integrated bioprocesses with energy and waste components require a thorough life-cycle and techno-economic analysis to assess the environmental impact and economic feasibility under different conditions of operation. The mechanisms and variations in microbial diversity with respect to the microbe
electrode interactions (anode/cathode) need to be further understood to enable improved CH4 production for industrial scale applications. These studies would benefit from a transfer of knowledge to large-scale operations with overall process understanding, while overcoming several technology related challenges. Symbiotic integration of multiple processes as a single unit can efficiently contribute to the cost economics and environmental sustainability along with specified products generation from the systems (Fig. 1.4).

Acknowledgments The Department of Biotechnology (DBT), Government of India, supported this research (BT/HRD/35/01/02/2018) in the form of Tata Innovation Fellowship to SVM. JSS acknowledges CSIR for providing research fellowship. The authors wish to thank CSIRIICT for supporting the research (Manuscript No. IICT/Pubs./2021/008).

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Talibi, M., Balachandran, R., Ladommatos, N., 2017. Influence of combusting methane-hydrogen mixtures on compression–ignition engine exhaust emissions and in-cylinder gas composition. Int. J. Hydrog. Energy. 42 (4), 2381
2396. Tan, L.C., Lin, R., Murphy, J.D., Lens, P.N.L., 2021. Granular activated carbon supplementation enhances anaerobic digestion of lipid-rich wastewaters. Renew. Energy. Available from: https://doi.org/10.1016/j.renene.2021.02.087. Tao, J., Qin, L., Liu, X., Li, B., Chen, J., You, J., Shen, Y., Chen, X., 2017. Effect of granular activated carbon on the aerobic granulation of sludge and its mechanism. Bioresour. Technol. 236, 60
67. Venkata Mohan, S., Butti, S.K., Amulya, K., Dahiya, S., Modestra, J.A., 2016a. Waste biorefinery: a new paradigm for a sustainable bioelectro economy. Trends Biotechnol. 34 (11), 852
855. Venkata Mohan, S., Chiranjeevi, P., Dahiya, S., Kumar, A.N., 2017. Waste derived bioeconomy in india: a perspective. N. Biotechnol. 40, 60
69. Venkata Mohan, S., Modestra, J.A., Amulya, K., Butti, S.K., Velvizhi, G., 2016b. A circular bioeconomy with biobased products from CO2 sequestration. Trends Biotechnol. 34 (6), 506
519. Venkata Mohan, S., Mohanakrishna, G., Goud, R.K., Sarma, P.N., 2009. Acidogenic fermentation of vegetable-based market waste to harness biohydrogen with simultaneous stabilization. Bioresour. Technol. 100 (12), 3061
3068. Venkata Mohan, S., Mohanakrishna, G., Sarma, P.N., 2008. Effect of anodic metabolic function on bioelectricity generation and substrate degradation in single chambered microbial fuel cell. Environ. Sci. Technol. 42 (21), 8088
8094. Venkata Mohan, S., Sarkar, O., 2017. Waste to biohydrogen: addressing sustainability with biorefinery. Energy Engineering. Springer, Singapore, pp. 29
37. Venkata Mohan, S., Velvizhi, G., Krishna, K.V., Babu, M.L., 2014. Microbial catalyzed electrochemical systems: a bio-factory with multi-facet applications. Bioresour. Technol. 165, 355
364. Vijayanand, C., Singaravelu, M., 2017. Installation of a biogas upgradation system. AgricINTERNATIONAL 4 (1), 12
15. Villano, M., Aulenta, F., Ciucci, C., Ferri, T., Giuliano, A., Majone, M., 2010. Bioelectrochemical reduction of CO2 to CH4 via direct and indirect extracellular electron transfer by a hydrogenophilic methanogenic culture. Bioresour. Technol. 101 (9), 3085
3090. Villano, M., Paiano, P., Palma, E., Miccheli, A., Majone, M., 2017. Electrochemically-driven fermentation of organic substrates with undefined mixed microbial cultures. ChemSusChem 10 (15), 3091
3097. Vrbova, V., Ciahotny, K., 2017. Upgrading biogas to biomethane using membrane separation. Energy Fuels 31 (9), 9393
9401. Xu, H., Wang, K., Holmes, D.E., 2014. Bioelectrochemical removal of carbon dioxide (CO2): an innovative method for biogas upgrading. Bioresour. Technol. 173, 392
398. Xu, J., Wu, H., Wang, Z., Qiao, Z., Zhao, S., Wang, J., 2018. Recent advances on the membrane processes for CO2 separation. Chin. J. Chem. Eng. 26 (11), 2280
2291. Xu, J., Zhang, Y., He, Y., Xie, L., 2020. Biogas upgrading. Adv. Bioenergy 5, 309
344. Yang, P., Tan, G.Y.A., Aslam, M., Kim, J., Lee, P.H., 2019. Metatranscriptomic evidence for classical and RuBisCO-mediated CO2 reduction to methane facilitated by direct interspecies electron transfer in a methanogenic system. Sci. Rep. 9 (1), 1
7. Yuan, B., Wu, X., Chen, Y., Huang, J., Luo, H., Deng, S., 2013. Adsorption of CO2, CH4, and N2 on ordered mesoporous carbon: approach for greenhouse gases capture and biogas upgrading. Environ. Sci. Technol. 47 (10), 5474
5480.

26

PART | I Introduction

Zhao, Z., Li, Y., Quan, X., Zhang, Y., 2017. Towards engineering application: Potential mechanism for enhancing anaerobic digestion of complex organic waste with different types of conductive materials. Water Res. 115, 266
277. Zhao, Z., Zhang, Y., Yu, Q., Ma, W., Sun, J., Quan, X., 2016. Enhanced decomposition of waste activated sludge via anodic oxidation for methane production and bioenergy recovery. Int. Biodeterior. Biodegrad. 106, 161
169. Zhen, G., Lu, X., Kobayashi, T., Su, L., Kumar, G., Bakonyi, P., et al., 2017. Continuous microcurrent stimulation to upgrade methanolic wastewater biodegradation and biomethane recovery in an upflow anaerobic sludge blanket (UASB) reactor. Chemosphere 180, 229
238. Zhou, K., Chaemchuen, S., Verpoort, F., 2017. Alternative materials in technologies for biogas upgrading via CO2 capture. Renew. Sustain. Energy Rev. 79, 1414
1441.

Chapter 2

Chemical absorption—amine absorption/stripping technology for biogas upgrading Alastair James Ward, Lu Feng and Henrik Bjarne Møller Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark

Chapter outline 2.1 Introduction 2.2 Process fundamentals 2.2.1 Amine chemistry 2.2.2 Amine selection 2.2.3 Process description and technology 2.2.4 Energy consumption 2.2.5 Operational problems and emissions 2.2.6 Economic considerations

2.1

29 31 31 32 34 42 43 45

2.3 Research and development directions 2.3.1 Novel liquid absorbents 2.3.2 Water-lean solvents/ nonaqueous amine solvents 2.3.3 Amine-functionalized solid sorbents 2.3.4 Process optimization 2.4 Conclusions and future perspectives References Further reading

47 47 48 48 49 50 50 55

Introduction

Chemical absorption is based on the CO2-reactive absorbents such as aqueous solutions of amine, or alkali aqueous solutions such as KOH, K2CO3, NaOH, Fe(OH)3, and FeCl2 (Lasocki et al., 2015). Amine absorption/stripping is the most common industrial process for CO2 capture and is a fairly well developed technology (Simmonds et al., 2003). Amine scrubbing using aqueous monoethanolamine (MEA) was first patented in 1930 for the removal of acid gases (CO2 and H2S) from natural gas, and is often referred to as first-generation amine scrubbing technology (Yuan and Rochelle, 2019). During chemical absorption, CO2 is absorbed in the liquid and reacts with the chemical substance at low temperature in the absorption column Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00003-9 © 2021 Elsevier Inc. All rights reserved.

29

30

PART | II Physiochemical upgrading systems

(Ryckebosch et al., 2011). Afterward, the chemical is regenerated with heat, usually as steam, in a desorption (stripping) column and CO2 can be released or recovered. It is necessary for the amine solution to be stripped and reused for economic and environmental reasons, as the CO2 emission during amine production is greater than that which can be absorbed (Puxty and Maeder, 2016). Two major advantages of amine absorption are that the process has very little CH4 loss (,0.1%) and the methane content in the scrubbed gas can be greater than 99% because the chemical solvent reacts selectively with CO2 (Awe et al., 2017). Advantages and disadvantages of the process are summarized in Table 2.1. In contrast to purely physical scrubbing, amine absorption proceeds at close to atmospheric pressure, and is easy to operate at low cost (Wellinger et al., 2013). H2S removal (often carried out in activated carbon filters) prior to amine scrubbing is highly recommended to prevent amine contamination, although some commercial units can cope with up to 300 ppm of H2S (Mun˜oz et al., 2015). Another advantage of using amine for biogas upgrading is the fact that there have been numerous experiences and extensive knowledge in both natural gas absorption-stripping processes (Kidnay et al., 2011) and CO2 capture from combustion flue gases (Dutcher et al., 2015). The use of amine absorption for biogas upgrading is in some ways more suitable than for scrubbing flue gases, as the latter contain relatively high concentrations of oxygen, which chemically attack amines. However, the use of biological sulfide removal systems prior to the chemical absorption process will also add a small amount of oxygen to the biogas. In addition, the relatively low CO2 partial pressure of flue gases leads to a higher energy requirement (Feron, 2016) than for biogas upgrading. In the biogas sector, the industrial application of amine CO2 removal has become one of the dominant technologies (Dawodu, 1994; Yu et al., 2012), only falling behind pressure swing absorption and water scrubbing in terms of popularity (Bauer et al., 2013b).

TABLE 2.1 Advantages and disadvantages of chemical absorption with amine (Ryckebosch et al., 2011). Advantages

Disadvantages

High efficiency ( . 99% CH4) Cheap operation Regenerative More CO2 dissolved per unit of volume (compared to water) Very low CH4 losses (,0.1%)

Expensive investment Heat required for regeneration Corrosion Decomposition and poisoning of the amines by O2 or other chemicals Precipitation of salts Foaming possible

Chemical absorption for biogas removal Chapter | 2

31

Limitations of amine solvents such as MEA are usually associated with a high energy requirement, significant amine degradation and loss by evaporation, corrosion problems, and foaming (Table 2.1) (Privalova et al., 2012). Most of the energy consumption is associated with heating of the aqueous amine solution from the absorption temperature to the desorption temperature and with the evaporation of solvent in the desorber column (Veneman et al., 2015). The desorption temperatures are in the range of 100
C130
C for amine regeneration in the stripper column but this varies with the choice of amine solvent (Sun et al., 2015). For instance, biogas upgrading using MEA requires more heat to regenerate the amine in the stripper column than methyl diethanolamine (MDEA) (Lange` et al., 2015). The development of improved solvents with lower regeneration energy requirements, therefore, can be identified as the highest priority research and development (R&D) objective for amine-based CO2 capture systems (Singh and Versteeg, 2008). Another major drawback associated with chemical absorption using amines is degradation through irreversible side reactions, mainly with CO2 (Gouedard et al., 2012). These reactions can lead to different problems with the process: solvent loss, formation of volatile compounds which are potentially hazardous for the environment, foaming, fouling, and corrosion (Islam et al., 2011). To solve these issues, significant research efforts have been dedicated to developing substituted amine-based sorbents with high CO2 capacity, low regeneration energy requirement, and reduced environmental impact (Privalova et al., 2013).

2.2 2.2.1

Process fundamentals Amine chemistry

The amines that have been most widely used for CO2 removal of acid gases are MDEA, diethanolamine (DEA), and MEA (Bauer et al., 2013a). However it has become common to use a mixture of MDEA and piperazine (PZ) called activated MDEA (aMDEA) (Kohl and Nielsen, 1997). It was released by BASF for licensing as aMDEA technology in 1982 (Comyns, 2014). Since the patent on the use of aMDEA has expired, the solvent can now be obtained from several amine suppliers such as Dow, Huntsman, and INEOS (Mokhatab and Poe, 2012). AMDEA has a significantly higher absorption capacity compared to MDEA alone. This is because secondary and primary amines such as PZ have high reaction rates for CO2 absorption and an ability to react the CO2 further with the MDEA (tertiary amine), which with its relatively low heat of reaction makes the regeneration process more efficient (Bishnoi and Rochelle, 2000; Zhang et al., 2001). The reactive nature of the aqueous amine solutions with CO2 is well known, and there has been considerable experimental and theoretical work done detailing the mechanisms and rates of these reactions (Santos, 2013).

32

PART | II Physiochemical upgrading systems

The principal reaction of interest between CO2 and a primary or secondary amine (in water) is typically considered to occur via the formation of a zwitterion and then to form carbamate (Laddha and Danckwerts, 1981): RR0 NH 1 CO2 x RR0 NH1 COO2 ðzwitterionÞ

ð2:1Þ

RR0 NH1 COO2 1 RR0 NH x RR0 NH1 COO2 ðcarbamateÞ 1 RR0 NH1 2 ð2:2Þ Combining these two reactions can be represented as (Penny and Ritter, 1983): 2RR0 NH 1 CO2 x R0 NH1 COO2 1 RR0 NH1 2

ð2:3Þ

RR0 1 NCOO2 1 H2 O x RR0 NH 1 HCO2 3

ð2:4Þ

The mechanisms involved in the absorption of CO2 by aqueous solutions of tertiary amines, such as MDEA, are somewhat different to primary and secondary amines as they do not react directly with CO2. In fact, they act as a base, catalyzing the hydration of CO2. Thus the reaction of interest in aqueous solutions of tertiary amines is (MacDowell et al., 2010; Santos, 2013): CO2 1 H2 O 1 R1 R2 R3 N"R1 R2 R3 NH1 1 HCO2 3

2.2.2

ð2:5Þ

Amine selection

Solvent selection is very important for the overall performance as it determines the CO2 capture loading and regeneration, solvent-induced equipment corrosion or environmental and public health impacts from emissions of solvents or byproducts. The chemical structures of primary, secondary, tertiary, and sterically hindered amines are shown in Fig. 2.1. The primary amine MEA is considered as the benchmark solvent to which alternative solvents must be compared (Kohl and Nielsen, 1997). The reactivity of amines to CO2 follows the order of primary, secondary, and tertiary amines. For example, the reaction constants with CO2 are 7000, 1200, and 3.5 m3/s per kmol for MEA, DEA, and MDEA, respectively, at 25
C (Bishnoi and Rochelle, 2002). Loading capacity is the amount of CO2 that the absorbent can contain, usually in mol CO2/mol amine. Primary amines have a low loading capacity: for example 7 molar MEA has a loading capacity of 0.5 mol CO2/mol amine, whereas other amines such as 7 M DEA are higher at 0.8 mol CO2/mol amine and 3.4 M MDEA, 0.58 mol CO2/mol amine (Li, 2015). In addition to the aforementioned compounds, sterically hindered amines, such as 2-amino2-methyl-1-propanol (AMP), have been proposed as the steric character reduces the stability of the formed carbamate. Thus, carbamate can undergo hydrolysis to form bicarbonate and in the meantime release free amine molecules for further reaction with CO2 and consequently enhance the CO2

Chemical absorption for biogas removal Chapter | 2

33

FIGURE 2.1 Amine molecular structures: (A) primary (MEA); (B) secondary (DEA); (C) tertiary (MDEA); and (D) sterically hindered (PZ).

equilibrium loading capacity up to 1.0 mol of CO2 per mol of amine, as high as that of a tertiary amine (Yu et al., 2012). PZ is a cyclic diamine (Fig. 2.1) which has many attractive properties. The presence of the methyl group significantly reduces the stability of the carbamate bond, resulting in the preferred formation of the bicarbonate, leading to the particularly high loading capacity of this solvent (Gouedard et al., 2012). Bishnoi and Rochelle (2000) studied the aqueous PZ system and showed that the rate constant of PZ with CO2 is an order of magnitude higher than that of conventional carbamate formers such as MEA. PZ has typical loading capacities around 0.76 (5 M) and 0.79 (8 M) (Li, 2015). When used in treating biogas to meet the tight CO2 specification of a liquid natural gas plant, the activity of CO2 absorption is too slow with pure MDEA, which must be enhanced with a promoter (Mokhatab et al., 2013). For instance, aMDEA is based on applying MDEA with PZ as an activator, combining the advantages of chemical and physical solvents. Compared to MEA, it has higher loading capacity, for example, 0.8 mol CO2/mol amine for a 2 M PZ 1 7 M MDEA mix (Li, 2015). The mix also has a lower energy requirement for regeneration. Since CO2 binds chemically to MDEA much less strongly than to MEA, therefore, the partial regeneration of the solvent can be carried out by simple flashing (physical characteristics). However, PZ is also a strong base which will compete with the other amine for the proton. The solution is to introduce steric hindrance which will not significantly affect the Brønsted basicity toward the smaller proton but it does reduce the interaction with the much larger CO2 molecule (Puxty and

34

PART | II Physiochemical upgrading systems

Maeder, 2016). This is why the mixture of PZ and AMP performs particularly well (AMP is a primary amine with a large enthalpy of protonation) (Seo and Hong, 2000). The character of this solution is more like a hybrid of being a strong chemical and a purely physical solvent. There are reports of a substantial reduction in energy requirements and a modest reduction in circulation rates for amine blends relative to the corresponding single amine systems of similar total amine concentration (Yang et al., 2008). For instance, there is a significant difference between the absorption capacity of the MDEA alone and the mixture of primary amines (PZ). The major part of the operational costs is caused by the solvent regeneration (up to 40 percent) (Santos, 2013).

2.2.3

Process description and technology

2.2.3.1 General process A diagram of a typical overall chemical absorption process is shown in Fig. 2.2. The process consists of an absorption stage, where CO2 in the biogas is dissolved into the amine liquid and a stripping stage where the amine solution is stripped of CO2 and returned to the absorber. The process is usually controlled by temperature swing, with a lower temperature in the absorber of 40
C65
C and a higher temperature in the stripper of 100150
C, although it is also possible to control the process through pressure swing (Puxty and Maeder, 2016). The temperature of the inlet absorbent is typically 20
C40
C but this increases down the absorption column due to the exothermic nature of the CO2 absorption reaction. Amine solution is added in excess to the expected CO2 concentration, typically 47 times more on a molecular basis, to prevent equilibrium constraints (Bauer et al., 2013b). The absorber operates at a low pressure of 12 bar. The upgraded biogas exits at the top of the absorber, whereas the CO2-rich amine leaving the absorber is heated by the stripper exit stream via a heat exchanger, then

FIGURE 2.2 Simplified process flow diagram of an amine scrubber.

Chemical absorption for biogas removal Chapter | 2

35

led into the top of the stripper column. A reboiler is located at the bottom of the stripper to apply the necessary heat for CO2 stripping and to generate steam to reduce the CO2 partial pressure. The stripper pressure is higher than the absorber, at 1.53 bar. The exiting CO2 and steam are cooled in a condenser, the condensate being returned to the stripper column (Bauer et al., 2013b). The lean amine stream is further cooled after the heat exchanger, before returning to the top of the absorption column.

2.2.3.2 Absorption and desorption columns Both the absorber and stripper are gasliquid contactors, consisting of vertical columns. In the absorber, the gas typically enters from the bottom and the liquid from the top, through a distribution apparatus where necessary to ensure an even spread of fluids across the width of the column. The countercurrent system allows the reduced CO2 partial pressure at the top of the absorption column to be exposed to freshly stripped absorbent, whereas the CO2-rich absorbent at the base of the column is still capable of removing some CO2 from the relatively higher CO2 partial pressure of gas entering at the base of the absorber column. Due to the excess of amine in the process, the driving force is the CO2 partial pressure in the gas. Increasing the rate of absorption allows for a smaller absorber column and therefore reduced capital expenditure. From a physical point of view, optimal absorption rates require the maximum interfacial area between the gas and liquid. The mass transfer between the fluids depends on diffusion, a slow process, the rate of which can be increased by turbulence to transport the gas molecules away from the interface area and into the bulk phase (Gruenewald and Radnjanski, 2016). As the basic amines have a relatively high pH, there is little risk of bacterial biofilm formation in the gasliquid contactors, which could potentially reduce flow in some designs. Therefore there are several suitable methods of achieving a high gasliquid interface area (Gruenewald and Radnjanski, 2016): 1. Packed columns. The liquid is spread thinly across solid surfaces in a continuous gas phase. The solid surfaces are either random or structured packing material. 2. Tray columns. Gas bubbles are dispersed into a continuous liquid phase through a vertical series of trays. 3. Bubble columns. A simple continuous bubble column. 4. Spray columns. The liquid is sprayed as small droplets through a continuous gas phase. 5. Membrane contactors. Porous hollow-fiber membranes separate the liquid and gas phases. Packed column reactors have a reasonable mass transfer rate and a low gasside pressure drop (Gruenewald and Radnjanski, 2016). They display flow

36

PART | II Physiochemical upgrading systems

characteristics that are very close to those of a plug flow reactor. The plug flow pattern assumes no back-mixing and a uniform velocity profile, although an increase in velocity may occur at the reactor wall due to point contact with the packing material and the reactor wall (Eigenberger, 1992). It can, however, be difficult to assure proper liquid distribution throughout a packed column, as preferential flow pathways may occur (Towler and Sinnot, 2013). Suitable random packing materials include plastic or stainless steel rings such as Raschig or Pall rings, although it is recommended that stainless steel rings be used in the stripper due to the high temperatures (Vo et al., 2018). The packing material can be either random or structured. Random packing often consists of steel rings, whereas structured packing consists of stacked corrugated and perforated sheets. There is evidence to suggest that structured packing performs better than random: Aroonwilas et al. (1999) compared CO2 absorbance in packed columns containing structured and random packing materials with either MEA or NaOH as solvents. The structured packing consisted of Gempak 4A, a corrugated material with a lanced and perforated surface, whereas the random packing was Intalox Metal Tower Packing (#25, #50, and #70) stainless steel rings. The results showed that the structured packing had a mass transfer rate that was double that of the best random packing (IMTP #25) tested. This was attributed to the increased packing surface area of 446 m2/m3 for Gempak 4 A compared to 230 m2/m3 for IMTP #25. However, it should be noted that the experiments were based on flue gas conditions with a low CO2 partial pressure of 1 kPa. Yeh et al. (2001), using a higher CO2 concentration of 15%, found BX Gauze structured packing to be better than Intalox saddles random packing in terms of absorption rate, % efficiency and % absorbent utilization, but Flexipac structured packing had only a slightly higher performance than the random packing. Fig. 2.3 shows examples of four types of random packings (RVT Process Equipment GmbH), with the respective technical data in Table 2.2.

(A)

(C)

(B)

(D)

FIGURE 2.3 Examples of random packings: (A) Raflux stoneware 25-3; (B) Hiflow polypropylene 15-7; (C) Hiflow polyethylene 25-7; and (D) Hiflow carbon steel 25-5.

Chemical absorption for biogas removal Chapter | 2

37

TABLE 2.2 Technical data of the random packings shown in Fig. 2.3. Type

Size

Surface area (m2/m3)

Void %

Raflux stoneware

25-3

220

73

Hiflow polypropylene

15-7

313

91

Hiflow polyethylene

25-7

214

92

Hiflow carbon steel

25-5

185

95

Tray columns have a higher gas-side pressure drop than packed columns but have improved mass transfer rates. However, they are less suitable for foaming liquids (Gruenewald and Radnjanski, 2016). They consist of a stack of perforated trays. The gas rises through the perforations and the liquid flows over a weir and down to the next lower tray via a downcomer at the edge of each tray. Tray columns are less compact than packed columns, as the mass transfer process occurs at discrete points (i.e., in the trays) rather than throughout the column, as is the case for packed columns. They exist as the three basic designs described below (Towler and Sinnot, 2013) and illustrated in Fig. 2.4AC. 1. Sieve trays (Fig. 2.4A) are the simplest designs, allowing the gas flowing upwards through perforations to bubble through the liquid lying on each tray surface. The liquid is prevented to some extent from falling through the perforations by the gas flow, although some weeping will occur which reduces efficiency. 2. A more complicated design is the bubble-cap tray (Fig. 2.4B), which has a tubular riser on the top side of each tray perforation, capped by an inverted cup. The gas bubbles out through slots in the side of the cap or serrations around the base of the cap. The bubble-cap tray maintains the same liquid level in each tray, regardless of gas flow rate, so the performance at low flow levels is good but they are prone to fouling. 3. Valve trays (Fig. 2.4C) use caps over the tray perforations, which are lifted by the gas pressure and the gas then bubbles into the liquid on the tray. This design forces the gas sideways through the liquid, which improves contact time and they are effective at low gas flow rates as the valves close under such circumstances. Bubble columns have the benefit of simple construction, high heat and mass transfer coefficients, and high removal efficiency (Chen et al., 2008). The column contains the liquid traveling downwards and the gas introduced through a sparger. The mass transfer rate depends on the bubble size distribution, which is in turn dependent on the sparger design, operating

38

PART | II Physiochemical upgrading systems

FIGURE 2.4 (A) Sieve tray gas contactor design including downcomer from tray above and overflow weir. (B) Bubble-cap tray design. Downcomers and overflow weirs (as detailed in A) are not shown. (C) Valve-cap tray design. Downcomers and overflow weirs (as detailed in A) are not shown.

Chemical absorption for biogas removal Chapter | 2

39

conditions, and physical properties of the gas and liquid phases (Gruenewald and Radnjanski, 2016). Chen et al. (2008), in an experiment using barium chloride as an absorbent, found mass transfer coefficients for bubble columns were lower than for random or structured packed columns. However, a scrubbing factor, φ, defined as moles of CO2 removed per mole of absorbent and per unit volume of the scrubber, was found to be considerably higher for the bubble column than the packed columns. Spray columns, like bubble columns, are simple in design. The liquid is sprayed in fine droplets through the gas. Spray columns offer low gas-side pressure drop but tend to have a lower degree of separation (Gruenewald and Radnjanski, 2016) and have not been widely studied for amine-CO2 removal. Membrane contactors have been developed to reduce absorbent volatilization (and subsequent loss) by choosing a membrane that presents a barrier to amines but allows a high CO2 flux (Bernhardsen et al., 2019). Membrane contactors have several advantages: The interfacial area is known and remains the same at high or low flow rates as the fluids remain independent of each other, making the performance easy to predict, they are easily scalable as they can be produced in modular form and they offer a very high interfacial area. However, the membrane is a barrier to mass transfer and can be subject to fouling (Gabelman and Hwang, 1999), although a study has shown that membrane contactors can have a higher mass transfer rate than packed columns (deMontigny et al., 2005). Wetting of the membrane also reduces mass transfer (Bernhardsen et al., 2019). The problem of membrane wetting is that the liquid absorbent floods the pores of a porous membrane, but the problem can be minimized with the use of nonporous thin-composite membranes (Ansaloni et al., 2017), such as a perfluoropolymer layer over a porous polypropylene layer (Ansaloni et al., 2019). As mentioned above, packed columns are generally more compact than tray designs and more resistant to corrosion but are less suitable than tray columns for situations with large variations in flow rate (Towler and Sinnot, 2013). However, flow of biogas from a particular plant is easily estimated and constant. Therefore packed columns are more common for amine biogas upgrading systems. The dimensioning methods for absorbers and strippers are the same. For random packing, Towler and Sinnot (2013) recommend that columns with a diameter of less than 30 cm should use a packing of ,25 mm, those of 3090 cm should use a packing of 2538 mm, and diameters .90 cm should use a packing of 5075 mm. Smaller packing is more expensive but has a greater mass transfer. A packing that is too large for the column diameter may suffer from poor liquid distribution. Specific surface areas of packing decrease with size, with values of 340480 m2/m3 for 1316 mm packing and 6685 m2/m3 for 7689 mm packing. Exact values for a specific packing should be obtained from the supplier.

40

PART | II Physiochemical upgrading systems

To scale column height, it is more common to use height units (Towler and Sinnot, 2013). The following equations are suitable for solute concentrations up to c. 10%, which is suitable for many loaded amines. The height of the packing (Z) can be expressed as transfer units, which will be of a certain height (HTU) and number (NTU). In terms of KG, the gas phase mass transfer coefficient: Z 5 HOG NOG

ð2:6Þ

where HOG is the height of a gas transfer unit: HOG 5

Gm KG aP

and NOG is the number of gas transfer units: ð y1 dy NOG 5 y2 y 2 y e

ð2:7Þ

ð2:8Þ

a 5 surface area of interface, P 5 total pressure, y1 5 molar fraction of CO2 in the gas at the bottom of the column, y2 5 molar fraction of CO2 in the gas at the top of the column, ye 5 mole fraction in the gas in equilibrium with the liquid at any point. The equations are similar in terms of KL, the liquid phase mass transfer coefficient: Z 5 HOL NOL

ð2:9Þ

HOL is the height of a liquid phase transfer unit: HOL 5

Lm KL aC1

NOL is the number of liquid phase transfer units: ð x1 dx NOL 5 x 2x e x2

ð2:10Þ

ð2:11Þ

a 5 surface area of interface, C1 5 total molar concentration, x1 5 molar fraction of CO2 in the liquid at the bottom of the column, x2 5 molar fraction of CO2 in the liquid at the top of the column, xe 5 mole fraction in the liquid in equilibrium with the gas at any point. Mass transfer coefficients for the relatively high CO2 partial pressures found in biogas are less well studied than for the lower partial pressures of flue gases. A selection of experimental KG values at PCO2 values more typical of biogas are listed in Table 2.3 (Wanderley et al., 2019). For scaling the column diameter, one must determine the maximum gas velocity. Higher gas velocities increase the resistance encountered by the liquid and therefore increase the pressure drop (ΔP/L) through the packing.

Chemical absorption for biogas removal Chapter | 2

41

TABLE 2.3 Overall mass transfer coefficients of a range of water-lean amines. 2MPZ (2-methylpiperazine), MDEA (N-methyl diethanolamine), SULF (Sulfolane), MEG (ethylene glycol), DMSO (dimethyl sulfoxide), NMP (N-methyl-2-pyrrolidinone). Amine

Amine conc.

CO2 loading α (mol/ mol N)

PCO2 (kPa)

Temp. (
C)

Overall mass transfer coefficient KG (mol/m2 s Pa)

2MPZ 1 SULF

2.5 M

0.40

20 6 2

40

1.9 3 1027

2MPZ 1 MEG

2.5 M

0.40

20 6 2

40

9.5 3 1028

2MPZ 1 DMSO

2.5 M

0.40

21 6 2

40

1.27 3 1027

2MPZ 1 NMP

2.5 M

0.39

34 6 3

40

7.8 3 1028

MDEA

3.5 M

0.35

22 6 9

40

1.6 3 1028

MDEA 1 SULF

3.5 M

0.20

30 6 7

40

4.7 3 1028

Pressure drop in the absorber should be lower than 100 mbar (Alix and Raynal, 2009). When gas velocity is too high, the column will become flooded with liquid and the increased pressure could damage the packing. The gas load factor (F-factor, FS) is calculated thus: pffiffiffiffiffiffi FS 5 ρG VSG ð2:12Þ where ρG is gas density and VSG is superficial gas velocity. The loading point of the column is defined as the point where gas velocity is sufficiently high to restrict the flow of liquid, above which the pressure drop increases at a faster rate. The flooding point is where the gas velocity is high enough to cause liquid hold up in the column (Fig. 2.5). Once the flooding point is determined for the design liquid flow rate, the dimensioning of the column can then be determined using the gas velocity between 50% and 80% of the flooding point. The throughput is also limited by the drip point density (dp) of the liquid distributor (per m2) and the flooding point. Sufficient drip points are recommended to ensure that this is not a limiting factor. According to Bauer et al. (2013b), the flooding point of the packing (F) can be determined by: sffiffiffiffiffiffiffiffiffiffiffiffi  ffi L ρG F5 ð2:13Þ G ρL where L is the liquid flow and G is a packing-specific function of the gas velocity. ρG and ρL are the densities of the gas and liquid, respectively.

PART | II Physiochemical upgrading systems

Pressure drop (ΔP/L)

42

Loading point Flooding point

FS FIGURE 2.5 Illustration of the loading point and flooding point of packing effect on pressure drop at increasing FS.

Pressure drop can be calculated using the Buchanan equation, as discussed in Brunazzi et al. (2009). However, curves illustrating ΔP/L versus FS for a range of L values can be sourced from the packing supplier. As an example of pressure drop at full scale, Alix and Raynal (2009) presented data for a flue gas CO2 absorption column of 1100 mm diameter with four packed beds, each of 4.25 m height using IMTP50 random stainless steel packing and MEA as an absorbent. At a liquid loading (QL) of between 20 m3/m2 per hour and FS values between 1 and 2 Pa0.5, ΔP/L was measured at c.23 mbar/m column height, which was found to be below the loading point.

2.2.4

Energy consumption

The amine absorption process has a low electrical energy requirement for pumping and minor tasks, yet has a much larger heat requirement, for the steam stripping, when compared to other biogas upgrading methods (Bauer et al., 2013b). Vo et al. (2018) modeled an electrical consumption of 0.023 kWh/m3 methane but a very high steam requirement for the stripper in excess of 2 kWh/m3 methane for a simulation of a process with c. 51% CH4 in the raw biogas and the stripper operating at 120
C and 1.5 bar. Bauer et al. (2013b) reported a much lower heat requirement of 0.55 kWh/m3 methane, yet a higher electrical consumption of approximately 0.12 kWh/m3 methane in a study averaging data from 22 manufacturers and users of amine biogas upgrading plants across Europe. Dixit (2015) provided a range of values between the previous two sources, with an electrical requirement of 0.060.15 kWh/m3 methane and a heat requirement of 0.50.8 kWh/m3 methane. Collet et al. (2017) collected information from a variety of sources for a life cycle assessment study using figures of 0.16 kWh/m3 methane for electrical consumption (including for compression of CO2) and 0.75 kWh/m3 methane heat consumption. Ideally, the temperature swing absorption process should have a low absorption temperature and high desorption temperature. The maximum

Chemical absorption for biogas removal Chapter | 2

43

temperature in the stripper is restricted by thermal degradation of the amine (Section 2.2.5.2). Li (2015) summarized the research into the performance parameters for a large range of amine solutions. The upper temperature tolerance limits ranged from 114
C [8 M (methylamino) propylamine (MAPA)] to 163
C (8 M PZ). For more common applications, 7 M MEA had a temperature tolerance up to 121
C, whereas a typical blended amine solution (2 M PZ 1 7 M MDEA) had a temperature tolerance up to 120
C. Richner et al. (2015) modeled the energy requirement of the reboiler, assuming isobaric conditions of the stripper, of a range of amines and mixtures of amines. The lowest tested reboiler energy requirement was 3.24 GJ/ton CO2 captured for 3 molar AMP with a lean amine (αlean) value of 0.03 mol CO2/ mol amine, a rich amine (αrich) value of 0.54 mol CO2/mol amine and a stripper base temperature of 104
C. The highest energy requirement was for MEA, at 3.97 GJ/ton CO2 with an αlean value of 0.22 mol CO2/mol amine, αrich value of 0.49 mol CO2/mol amine and a stripper base temperature of 116
C. The high heat requirement of the stripping process means there is a lot of low-grade heat available. Bauer et al. (2013b) claim that 80% of the heat can be reused for heating biogas reactors, for example.

2.2.5

Operational problems and emissions

2.2.5.1 Amine losses There is a risk of loss of amines as entrained droplets, mist, or through volatilization, with MEA being more volatile than MDEA. There can also be losses of amines through the fluid connections, such as pipe joints and heat exchangers (Bauer et al., 2013b). Amine loss into the gas phase is generally greater from the absorber than from the stripper. The amine treatment of biogas differs from flue gas treatment as both the methane and carbon dioxide gas outlet streams can be utilized and the volatilized amines will not be released into the environment. However, the contamination of the gas streams is a problem, as is the loss of absorbent. Vo et al. (2018) estimated an evaporative loss of 205 kg of MEA (around 6%) per year for a plant with an annual biomethane production of 3.44 million m3. Washing with water at the top of the columns is usually employed to reduce losses. The amine dissolves in the wash water and this is returned to the column. There can be several wash stages stacked one above the other, the solution runs down from a higher stage to the stage below, often of a bubble-cap tray design. Amine dissolved in the wash water will increase in concentration, going down from a higher wash stage to a lower stage and thus the washing operates in counter current. Data suggest that the amine concentration in the exit gas is around 0.010.05 ppm (Svendsen et al., 2011). The use of thin-composite membrane amine absorbers can reduce volatilized amine losses (Ansaloni et al., 2019).

44

PART | II Physiochemical upgrading systems

2.2.5.2 Degradation of absorbent Amines can also suffer from degradation during the process, although this is more of a problem for flue gases due to the presence of higher concentrations of oxygen (Reynolds et al., 2016) than is typically found in biogas. Thermal degradation in the presence of CO2 mainly occurs in the stripper and is possible if the amine contacts a surface in excess of 175
C. MEA thermally degrades to carbamate and then through oxazolidin-2-one (OZD) to further degradation products. DEA can degrade to MEA and therefore degrade further as described for MEA as well as through other pathways. Oxidative degradation of amines forms a wide variety of compounds. These include carboxylic acids, which cause corrosion and fouling and lead to the formation of heat-stable salts (Gouedard et al., 2012). Amine oxidation and the formation of heat-stable salts are considered the most rapid degradation reactions. Svendsen et al. (2011) suggest that the degradation products can be classified as below: G G G

Volatile (ammonia, aldehydes, etc.); Low volatility (volatility lower than ethanolamine); Nonvolatile (typically heat-stable salts, organic acids, etc.).

The degradation products must be removed continuously but the volatiles are often water-soluble and so will be collected in the washing stages until the water reaches saturation. Bleeding from the top wash can remove these with minimum loss of amines due to the low concentration at the top wash.

2.2.5.3 Methane losses Losses of methane are an important issue for an upgrading process, through loss of product and because methane is a potent greenhouse gas. Kvist and Aryal (2019) compared methane losses from three different upgrading technologies in nine biogas plants and found the amine scrubbing method to have a considerably lower methane loss (0.05%) than membrane separation (0.52%) or water scrubbing (1.52%). A 0.05% methane loss has also been reported by the Danish Gas Technology Center (Dansk Gasteknisk Center, 2018). Bauer et al. (2013b) reported that amine scrubbers have a maximum methane slip of 0.1%. 2.2.5.4 Foaming Foaming can be a problem with amine absorption processes. It is mainly caused by hydrocarbons such as oil residues in new equipment (Bauer et al., 2013b) or from corrosion inhibitors (Keewan et al., 2018) or suspended solids. The problem manifests itself as a high-pressure differential across either the absorption or desorption column, amine carry-over between columns, loss of amine solution, off-spec gas products, or poorly stripped solvent (Thitakamol et al., 2009; Bauer et al., 2013b). It is possible to treat

Chemical absorption for biogas removal Chapter | 2

45

foaming with commercial antifoaming agents. In an experimental setup using MEA, Thitakamol et al. (2009) found that foam height increased with gas flow rate, solution volume, CO2 loading, and MEA concentration, but decreased with temperature. The authors further developed a model constructed by Pilon et al. (2001) to predict pneumatic steady-state foam heights.

2.2.6

Economic considerations

When comparing the investment costs associated with chemical absorption biogas upgrading with other technologies, the chemical method had higher costs for smaller installations but all upgrading methods had very similar investment costs for large installations with raw biogas flow rates of more than 1500 Nm3/h (Bauer et al., 2013a). Specific investment costs follow a typical economy-of-scale curve, with values from Sweden (Bauer et al., 2013b), Denmark (Dansk Gasteknisk Center, 2018), and Ireland (Vo et al., 2018) tabulated in Table 2.4. The Bauer et al. (2013b) and Vo et al. (2018) values have been normalized for inflation to the year 2018 using Eq. (2.14) (Brown and Brown, 2013).   Ic Cp;c 5 Cp;p ð2:14Þ Ip where Cp,c is the adjusted cost for the year, Cp,p is the known cost in a previous year, Ic is the cost index for the year, and Ip is the cost index for the previous year. The data from Bauer et al. (2013b) were for 2013 prices, whereas Vo et al. (2018) published 2016 values.

TABLE 2.4 Approximate specific plant investment costs. Raw biogas flow rate (Nm3/h)

Specific investment costs (h/Nm3/h)

References

600

3521

Bauer et al. (2013b)

813

2191

Vo et al. (2018)

833

32013917

Dansk Gasteknisk Center (2018)

900

2708

Bauer et al. (2013b)

1800

1733

Bauer et al. (2013b)

2500

15131595

Dansk Gasteknisk Center (2018)

46

PART | II Physiochemical upgrading systems

TABLE 2.5 Costs of amines (Santos, 2013). Amine

Cost (h/L)

Diethylamine

66.40

MEA

30.50

PZ

68.70

EDA

31.70

MDEA

51.60

DEA

25.70

The values for the Danish Gas Technology Center (Dansk Gasteknisk Center, 2018), DGC, have been converted to Euros from Danish Kroner at an exchange rate of h1 5 DKK 7.47. The raw biogas flow rates have been back-calculated from the biomethane flow rates (500 and 1500 Nm3/h) using a biogas methane composition of 60%. The data have also been adjusted to remove a 6% interest rate over a 10-year loan period (Dansk Gasteknisk Center, 2018). When compared to other biogas upgrading technologies, chemical absorption has a relatively low electricity consumption but a high heat consumption. Therefore operating costs in part depend on the cost of electricity and available heat energy; typical energy consumption was discussed in Section 2.2.4. The loss of amine (Section 2.2.5.1) also adds to the operating costs. Santos (2013) listed the costs (in 2013) of a variety of amines (Table 2.5). Operating costs of amine absorption plants in Denmark in 2018 are estimated to be between 0.04 and 0.06 h/Nm3 methane for a plant of 833 m3/h biogas flow rate (500 m3/h methane flow rate) and between 0.025 and 0.05 h/Nm3 methane for a plant of 2500 m3/h biogas flow rate (1500 m3/h methane flow rate) (Dansk Gasteknisk Center, 2018). An overall operating cost of 1.97 million euros per year for a complete biogas plant and amine absorption upgrading plant, with annual biomethane production of 3.44 million m3, has been suggested. This operating cost was broken down into 38% for utilities, 35% raw materials, 15% labor, 10% facility-dependent (maintenance and miscellaneous costs), and 2% laboratory costs, including quality control/ assurance (Vo et al., 2018). Another parameter that affects the economy is the expected downtime of the plant. The report by DGC (Dansk Gasteknisk Center, 2018) gives a standard downtime of 2% of total, which was the same for all suppliers of amine, water adsorption, and membrane technologies for biogas CO2 scrubbing that were examined in the report.

Chemical absorption for biogas removal Chapter | 2

2.3

47

Research and development directions

Currently, amine scrubbing is one of the most widely used technologies for biogas upgrading but the technology is still developing. CO2 capture research tends to be dominated by amine-based technologies for postcombustion flue gas CO2 capture, which are well-known for the reversible reactions with CO2 (Dutcher et al., 2015). This provides a wealth of experience useful for developing advanced amine or amine-based technologies for biogas upgrading. It was shown that aMDEA systems are still the most popular choice economically. However, aMDEA CO2 capture technology has several drawbacks calling for research into new solutions, which are cheaper, more efficient, and environmentally friendly. The extensive energy loss due to regenerating the solvent pushes research toward the development of substitute sorbents with high adsorption capacity, low regenerating energy requirement, high CO2 selectivity, durability, and relatively fast kinetics of sorption and desorption (Samanta et al., 2012). In addition, reducing latent and sensible heat consumption through optimized process design is another active research area. In the following sections, there are several potential directions or applications which might be promising for developing commercial-scale aminebased biogas upgrading technologies.

2.3.1

Novel liquid absorbents

In recent years, some amino acid salt solutions have been proposed as an alternative to the conventional amine absorbents, which can be competitive in terms of thermal stability, selectivity, and price (Mun˜oz et al., 2009). Amino acids can replace alkanolamines for CO2 capture due to the presence of the amino group (van Holst et al., 2009). Some amino acid solutions (AASs) have been demonstrated to have affinity to CO2 at least equaling that of alkolamines due to identical amino functional groups, while having lower absorbent losses and better resistance to oxidative degradation (Song et al., 2012). AAS selectivity between CO2 and CH4 has been shown to be double that of MEA (Simons et al., 2010). AAS is also a “green” solvent as it is biodegradable and thus has a lower environmental impact during disposal (Yan et al., 2015). AASs are reaction products between amino acids and alkaline substances, which are prepared using inorganic strong alkali, organic amine, addition of activators, or ionic liquids (Zhang et al., 2018). Song et al. (2012) evaluated the CO2 absorption characteristics of aqueous amino acid salt solutions by screening 16 amino acids and some blends with PZ. The results showed that the alanine, serine, and α-aminobutyric acid salts had relatively fast initial rates of absorption and desorption, resulting in high net cyclic capacity.

48

PART | II Physiochemical upgrading systems

Commercially, Siemens have developed the PostCap AAS system that is claimed to be almost emission free (Kuettel et al., 2013). In addition, different AASs blended with other liquids to form novel absorbents have been reported (Zhang et al., 2018). In the solvent blends, AASs with higher CO2 loading can act as a promoter to enhance the overall absorption performance of blended solutions.

2.3.2

Water-lean solvents/nonaqueous amine solvents

Water-lean or nonaqueous amine solvents have been developed with comparable selectivity to water-based solvents but as the heat capacity of the organic solvents is usually lower than water, lower stripping temperatures are required as well as reduced corrosion of the equipment (Heldebrant et al., 2017). The first water-lean solvents replaced water with organic co-solvent with the initial purpose of enhancing the CO2 in tertiary alkanolamines using blends such as N-MDEA in methanol, TEA in alcohols, or DEA in ethylene glycol (Henni and Mather, 1995; Oyevaar et al., 1989; Sada et al., 1989). Recent advancements by several research groups have led to the development of nonaqueous amine solvents. For instance, Guo et al. (2019) investigated the absorptiondesorption performance of CO2 into several blends of MEA or DEA with glycol ethers [2-methoxyethanol (2ME) and 2ethoxyethanol (2EE)] as nonaqueous solvents while comparable absorption capacity with aqueous 5.0 M MEA and higher desorption efficiency resulting in a larger cyclic capacity were observed. Heldebrant et al. (2017) summarized a variety of water-lean solvents which can potentially be utilized for postcombustion CO2 capture. In their review, the authors concluded that water-lean solvents have acceptable water tolerance and faster CO2 mass transfer than aqueous solvents, although at a greater cost and higher viscosity. It was expected that large-scale testing would take place within a few years.

2.3.3

Amine-functionalized solid sorbents

The use of supported amine sorbents to remove CO2, H2S, and H2O from methane-rich biogas is a relatively new development proposed originally for postcombustion CO2 capture from flue gas. The technology has developed ¨ nveren et al., 2017). from the aqueous amines that have been used widely (U According to Samanta et al. (2012), supported amine sorbents are placed in three classes: Class 1: Physically loading monomeric or polymeric amine species onto a porous support, such as silica by impregnation; Class 2: Covalent tethering of amine-containing silane to a solid support such as porous silica amine, mainly amine-containing silane;

Chemical absorption for biogas removal Chapter | 2

49

Class 3: Amino polymers are polymerized onto porous supports in situ. This is a hybrid of classes 1 and 2. Compared to regular amine scrubbing processes, the advantages of applying solid amine for biogas upgrading were summarized by Sutanto et al. (2017) as: (1) reduction of the heat energy required for regeneration as the heat capacity of solid sorbent is normally lower than liquid and extra heat for evaporation of water can be avoided; (2) reduction of emissions from amine degradation and, therefore, minimize the risk of corrosion because the amines are immobilized into the sorbents; (3) simpler operation for small-scale systems as a gassolid fixed bed would replace gas/liquid circulation; (4) various materials can be selected and applied as amine-supported absorbent including silica (Min et al., 2017), alumina (Chaikittisilp et al., 2011), carbon (Wang et al., 2015), resin (Chen et al., 2013), metal organic frameworks (Lin et al., 2013), and treated porous materials (Vilarrasa-Garc´ıa et al., 2017).

2.3.4

Process optimization

There have been a number of modifications to the process, particularly to reduce latent and sensible heat demands. Jin et al. (2018) proposed four scenarios of optimization of a flue gas treatment process. The consecutive scenarios added a further process modification to the previous scenario. First, a pre-concentrating membrane, which increased CO2 concentration in the permeate stream to be added at the absorber base, whereas the lower CO2 concentration stream was fed to an intermediate height of the absorber. This gave a 2.5% saving in reboiler duty. Adding an intercooler to the absorber increased the reboiler saving to 8.7%. The next stage involved adding a rich-split process. This recovered heat in the steam generated in the stripper and gave a dramatic 27.1% total reboiler saving. Finally, an air stripper was used to strip the lean amine to create extra-lean solvent, with a total reboiler saving of 28.1%. However, the final modification will reduce the amount of CO2 that can be isolated and potentially sold. Li et al. (2016) also investigated absorber intercooling, which reduced regeneration duty from 3.6 to 3.55 MJ/kg CO2 and a 25% reduction in absorber packing height. The authors also proposed stripper interheating, using heat in the hot lean amine to reduce the sensible heat requirement of the reboiler by 6.7%. The richsplit process was modified to also use heat from the CO2 product stream condenser. The traditional rich-split process reduced reboiler duty by 8.3% and the modified version by 10%. Membrane evaporators or condensers have been used to regenerate the absorbent. In a membrane evaporator, the hot rich amine stream from the stripper allows the permeate to return to the column as stripping steam (Ma et al., 2017). Membrane contactors for absorption and stripping have many advantages over more traditional packed columns, as described in Section 2.2.3.2 (Gabelman and Hwang, 1999).

50

PART | II Physiochemical upgrading systems

Further developments to reduce heat consumption include stripping under vacuum at temperatures below 100
C. The heat demand is therefore for lowgrade heat rather than steam in the conventional stripping column. Ahn et al. (2013) combined vacuum stripping with membranes at temperatures between 25
C and 80
C. CO2 flux (with MEA as the absorbent) increased from 31 mol/m2 h to 92 mol/m2h with pressure decreasing from 34.7 kPa to 8.0 kPa at 80
C. DEA and TEA had CO2 fluxes over 120 mol/m2h at the lowest pressure and highest temperature investigated. However, temperature still plays an important role, as at 8.0 kPa and 25
C, CO2 flux (for MEA) was considerably lower at only c. 10 mol/m2 h. There is an additional requirement for a vacuum pump, which will increase electrical consumption by approximately 0.05 kWh/Nm3 (Bauer et al., 2013b) as well as increasing capital expenditure. An added advantage of a lower stripping temperature is reduced thermal degradation of the absorbent.

2.4

Conclusions and future perspectives

Chemical absorption of CO2 in biogas is a particularly promising technology due to the high quality of the scrubbed gas and low methane losses. The low electrical energy consumption is also an advantage, whereas the high heat energy consumption is a major disadvantage. This has directed research toward finding absorbents that have a lower heat of desorption while maintaining the specificity and affinity of the more traditional amine solutions with reduced environmental impacts. This includes AASs, water-lean/nonaqueous solvents or amine-functionalized solid sorbents. Heat energy consumption can also be reduced by improving the physical process, such as splitting the rich amine stream, increased use of heat exchangers to heat/cool the amine where necessary and vacuum stripping. Using membranes instead of packed bed or tray columns can reduce the loss of absorbent through volatilization. The application of some or all of the above improvements should ensure that chemical absorption remains a dominant technology for CO2 removal from biogas.

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Ma, Q., Fang, M., Wang, T., Yu, H., Feron, P.H.M., 2017. Membrane evaporation for energy saving in CO2 chemical absorption process using a polybenzimidazole film: mass and heat transfer. Energy Fuels 31 (10), 1109111098. MacDowell, N., Florin, N., Buchard, A., Hallett, J., Galindo, A., Jackson, G., et al., 2010. An overview of CO2 capture technologies. Energy Environ. Sci. 3 (11), 16451669. Min, K., Choi, W., Choi, M., 2017. Macroporous silica with thick framework for steamstable and high-performance poly (ethyleneimine)/silica CO2 adsorbent. ChemSusChem 10 (11), 25182526. Mokhatab, S., Mak, J.Y., Valappil, J.V., Wood, D.A., 2013. Handbook of Liquefied Natural Gas. Gulf Professional Publishing. Mokhatab, S., Poe, W.A., 2012. Natural Gas Sweetening, Handbook of Natural Gas Transmission and Processing, 2nd ed. Gulf Professional Publishing, pp. 253290. Mun˜oz, D.M., Portugal, A.F., Lozano, A.E., Jose´, G., de Abajo, J., 2009. New liquid absorbents for the removal of CO2 from gas mixtures. Energy Environ. Sci. 2 (8), 883891. Mun˜oz, R., Meier, L., Diaz, I., Jeison, D., 2015. A review on the state-of-the-art of physical/ chemical and biological technologies for biogas upgrading. Rev. Environ. Sci. Bio/Technol. 14 (4), 727759. Oyevaar, M.H., Fontein, H.J., Westerterp, K.R., 1989. Equilibria of carbon dioxide in solutions of diethanolamine in aqueous ethylene glycol at 298 K. J. Chem. Eng. Data 34 (4), 405408. Penny, D.E., Ritter, T.J., 1983. Kinetic study of the reaction between carbon dioxide and primary amines. J. Chem. Society, Faraday Trans. 1: Phys. Chem. Condens. Phases 79 (9), 21032109. Pilon, L., Federov, A.G., Raymond, V., 2001. Steady-state thickness of liquid-gas foams. J. Colloid Interface Sci. 242 (2), 425436. Privalova, E.I., Ma¨ki-Arvela, P., Murzin, D.Y., Mikkhola, J.-P., 2012. Capturing CO2: conventional versus ionic-liquid based technologies. Russian Chem. Rev. 81 (5), 435. Privalova, E., Rasi, S., Ma¨ki-Arvela, P., Era¨nen, K., Rintala, J., Murzin, D.Y., et al., 2013. CO2 capture from biogas: absorbent selection. RSC Adv. 3 (9), 29792994. Puxty, G., Maeder, M., 2016. Chapter 2, The fundamentals of post-combustion capture. Absorption-Based Post-Combustion Capture of Carbon Dioxide. Woodhead Publishing. Reynolds, A.J., Verheyen, T.V., Meuleman, E., 2016. Chapter 16, Degradation of amine-based solvents. Absorption-Based Post-Combustion Capture of Carbon Dioxide. Woodhead Publishing. Richner, G., Puxty, G., Carnal, A., Conway, W., Maeder, M., Pearson, P., 2015. Thermokinetic properties and performance evaluation of benzylamine-based solvents for CO2 capture. Chem. Eng. J. 264, 230240. Available from: https://doi.org/10.1016/j.cej.2014.11.067. Ryckebosch, E., Drouillon, M., Vervaeren, H., 2011. Techniques for transformation of biogas to biomethane. Biomass Bioenergy 35 (5), 16331645. Sada, E., Kumazawa, H., Ikehara, Y., Han, Z.Q., 1989. Chemical kinetics of the reaction of carbon dioxide with triethanolamine in non-aqueous solvents. Chem. Eng. J. 40 (1), 712. Samanta, A., Zhao, A., Shimizu, G.K.H., Sarkar, P., Gupta, R., 2012. Post-combustion CO2 capture using solid sorbents: a review. Ind. Eng. Chem. Res. 51 (4), 14381463. Santos, S.P., 2013. Comparative study of amine solutions used in CO2 absorption/desorption cycles, Instituto Superior de Engenharia de Lisboa. Seo, D.J., Hong, W.H., 2000. Effect of piperazine on the kinetics of carbon dioxide with aqueous solutions of 2-amino-2-methyl-1-propanol. Ind. Eng. Chem. Res. 39 (6), 20622067. Simmonds, M., Hurst, P., Wilkinson, M.B., Watt, C., Roberts, C.A., 2003. A study of very large scale post combustion CO2 capture at a refining & petrochemical complex. In: Gale, J., Kaya, Y. (Eds.), Greenhouse Gas Control Technologies - 6th International Conference. Pergamon, Oxford, pp. 3944.

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Simons, K., Nijmeijer, K., Mengers, H., Brilman, W., Wessling, M., 2010. Highly selective amino acid salt solutions as absorption liquid for CO2 capture in gasliquid membrane contactors. ChemSusChem 3 (8), 939947. Singh, P., Versteeg, G.F., 2008. Structure and activity relationships for CO2 regeneration from aqueous amine-based absorbents. Process Saf. Environ. Prot. 86 (5), 347359. Song, H.-J., Park, S., Kim, H., Gaur, A., Park, J.-W., Lee, S.-J., 2012. Carbon dioxide absorption characteristics of aqueous amino acid salt solutions. Int. J. Greenh. Gas Control. 11, 6472. Sun, Q., Li, H., Yan, J., Liu, L., Yu, Z., Yu, X., 2015. Selection of appropriate biogas upgrading technology-a review of biogas cleaning, upgrading and utilisation. Renew. Sustain. Energy Rev. 51, 521532. Sutanto, S., Dijkstra, J.W., Pieterse, J.A.Z., Boon, J., Hauwert, P., Brilman, D.W.F., 2017. CO2 removal from biogas with supported amine sorbents: first technical evaluation based on experimental data. Sep. Purif. Technol. 184, 1225. Svendsen, H.F., Hessen, E.T., Mejdell, T., 2011. Carbon dioxide capture by absorption, challenges and possibilities. Chem. Eng. J. 171, 718724. Available from: https://doi.org/ 10.1016/j.cej.2011.01.014. Thitakamol, B., Veawab, A., Aroonwilas, A., 2009. Foaming in amine-based CO2 capture process: experiment, modeling and simulation. Energy Proc. 1, 13811386. Towler, G., Sinnot, R., 2013. Chapter 17 Separation columns (distillation, absorption and extraction). Chemical Engineering Design: Principles, Practice and Economics of Plant and Process Design. 2nd edition. pp. 807935. ¨ ., Sarıo˘glan, S., ¨ nveren, E.E., Monkul, B.O U ¸ Karademir, N., Alper, E., 2017. Solid amine sorbents for CO2 capture by chemical adsorption: a review. Petroleum 3 (1), 3750. van Holst, J., Versteeg, G., Brilman, D.W.F., Hogendoorn, J., 2009. Kinetic study of CO2 with various amino acid salts in aqueous solution. Chem. Eng. Sci. 64 (1), 5968. Veneman, R., Frigka, N., Zhao, W., Li, Z., Kersten, S., Brilman, W., 2015. Adsorption of H2O and CO2 on supported amine sorbents. Int. J. Greenh. Gas Control 41, 268275. Vilarrasa-Garc´ıa, E., Cecilia, J., Bastos-Neto, M., Cavalcante Jr, C., Azevedo, D., Rodr´ıguezCastello´n, E., 2017. Microwave-assisted nitric acid treatment of sepiolite and functionalization with polyethylenimine applied to CO2 capture and CO2/N2 separation. Appl. Surf. Sci. 410, 315325. Vo, T.T.Q., Wall, D.M., Ring, D., Rajendran, K., Murphy, J.D., 2018. Techno-economic analysis of biogas upgrading via amine scrubber, carbon capture and ex-situ methanation. Appl. Energy 212, 11911202. Wanderley, R.R., Yuan, Y., Rochelle, G.T., Knuutila, H., 2019. CO2 solubility and mass transfer in water-lean solvents. Chem. Eng. Sci. 202, 403416. Wang, J., Huang, H., Wang, M., Yao, L., Qiao, W., Long, D., et al., 2015. Direct capture of low-concentration CO2 on mesoporous carbon-supported solid amine adsorbents at ambient temperature. Ind. Eng. Chem. Res. 54 (19), 53195327. Wellinger, A., Murphy, J.D., Baxter, D., 2013. The Biogas Handbook: Science, Production and Applications. Elsevier. Yan, S., He, Q., Zhao, S., Zhai, H., Cao, M., Ai, P., 2015. CO2 removal from biogas by using green amino acid salts: performance evaluation. Fuel Process. Technol. 129, 203212. Yang, H., Xu, Z., Fan, M., Gupta, R., Slimane, R.B., Bland, A.E., et al., 2008. Progress in carbon dioxide separation and capture: a review. J. Environ. Sci. 20 (1), 1427. Yeh, J.T., Pennline, H.W., Resnik, K.P., 2001. Study of CO2 absorption and desorption in a packed column. Energy Fuels 15 (2), 274278.

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Yu, C.-H., Huang, C.-H., Tan, C.-S., 2012. A review of CO2 capture by absorption and adsorption. Aerosol Air Qual. Res. 12 (5), 745769. Yuan, Y., Rochelle, G.T., 2019. CO2 absorption rate and capacity of semi-aqueous piperazine for CO2 capture. Int. J. Greenh. Gas Control 85, 182186. Zhang, X., Zhang, C.-F., Qin, S.-J., Zheng, Z.-S., 2001. A kinetics study on the absorption of carbon dioxide into a mixed aqueous solution of methyldiethanolamine and piperazine. Ind. Eng. Chem. Res. 40 (17), 37853791. Zhang, Z., Li, Y., Zhang, W., Wang, J., Soltanian, M.R., Olabi, A.G., 2018. Effectiveness of amino acid salt solutions in capturing CO2: a review. Renew. Sustain. Energy Rev. 98, 179188.

Further reading Kaparaju P., 2013. Biogas upgrading scenarios in Europe  status & prospects. In: International Workshop on Promotion of Biogas Upgrading and Bottling in India and EU August 2224, 2013. IIT Delhi. Bauer, Fredric, Hulteberg, Christian, Persson, Tobias, Tamm, Daniel, 2013. Biogas upgrading  Review of commercial technologies. Svensk Gastekniskt Center AB.

Chapter 3

Water scrubbing for biogas upgrading: developments and innovations Valerio Paolini, Patrizio Tratzi, Marco Torre, Laura Tomassetti, Marco Segreto and Francesco Petracchini National Research Council of Italy, Institute of Atmospheric Pollution Research, Monterotondo, Italy

Chapter outline 3.1 Introduction 57 3.2 Absorption methodologies 58 3.2.1 Absorption in water (water scrubbing) 58 3.2.2 Absorption in NaOH solutions (alkaline scrubbing) 60 3.2.3 Absorption in K2CO3 solutions (hot potassium carbonate) 61 3.3 Absorption configurations 62

3.1

3.3.1 Packed column reactors 3.3.2 Hollow fiber membrane contactors 3.4 Chemical promoters in water absorption 3.5 Energy consumption 3.6 Methane slip and efficiency 3.7 Conclusions References

62 63 64 66 67 68 68

Introduction

There are many well-known and tested technologies for CO2 removal from a gas stream, and several of these have been already applied to biogas upgrading with good results (Hoyer et al., 2016). Historically, water scrubbing for the upgrading of biogas to biomethane was already considered the most economically viable solution that was commercially available in 1978 (Chen et al., 1980), due to its simplicity. This idea remained valid in the 1980s (Sweeten et al., 1981), when the concept of combining water scrubbing with bioscrubbing was considered, as microorganisms would already grow on water scrubbing units, acting not dissimilarly to biological filters and needing similar design requirements (Wheatley, 1984). While in the 1990s an increase in studies pertaining to the upgrading of biogas may be seen, water

Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00001-5 © 2021 Elsevier Inc. All rights reserved.

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scrubbing and alkaline scrubbing remain the default options for the removal of CO2 and H2S from biogas, as regeneration of the solution in a stripping column keeps costs low and sustainable (Strevett et al., 1995; Shiga et al., 1998). During the first years of the 2000s the trend regarding the number of scientific papers on biogas upgrading kept growing together with the number of citations. At the same time, economic viability and simplicity of water scrubbing were still considered unsurpassed, together with satisfactory efficiency and the possibility of coupling CO2 removal with biomethane compression, as the interest in employing sustainable liquified petroleum gas (LPG) from biomass for transportation grew (Petersson and Wellinger, 2009; Tippayawong and Thanompongchart, 2010). Today, 379 biogas upgrading plants are active in Europe (Parisi et al., 2020); water scrubbing is employed in more than 30% of upgrading plants in Europe, confirming the maturity and continued use of this technology (Prussi et al., 2019). Recent research is focused on novel approaches to water scrubbing aided by alkaline chemicals (Paolini et al., 2019) on companion technologies to develop useful byproducts or environmental services together with biomethane (Batlle-Vilanova et al., 2019), and on the environmental evaluation of biogas upgrading technologies (Lombardi and Francini, 2020). This chapter analyzes the state of the art of biogas upgrading through water scrubbing based on recent studies and evaluates its advantages and disadvantages in terms of absorption methodologies and operative parameters such as methane slip, energy consumption, and efficiency. Water scrubbing is a mature solution which is less sensible to biogas impurities than most other upgrading technologies and as such is widely used nowadays in more than 40% of upgrading plants (Munoz et al., 2015).

3.2 3.2.1

Absorption methodologies Absorption in water (water scrubbing)

In water scrubbing, CO2 molecules are sorbed by means of weak molecular forces into the liquid matrix, and it is usually performed at low temperatures and high pressure to further increase CO2 solubility. Water is then directed to a desorption tower, where it can be regenerated to be then recirculated by adding atmospheric air, lowering the pressure and raising the temperature, thus reducing CO2 solubility in water and releasing it into the atmosphere. The volume of water needed to remove a certain amount of CO2 from the biogas stream depends on the physical structure and packing of the water scrubbing tower: as the path the gas needs to travel through and the interactions between gas and liquid molecules grow, lower water volumes are needed to capture most CO2 molecules. In order to reach very low CO2 concentrations, on the other hand, the water volume needed grows as the partial pressure of the remaining carbon dioxide lowers (Bauer et al., 2013).

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This type of absorption is very selective as the solubility of carbon dioxide and methane in water are largely different, as shown by Henry’s constants in Table 3.1. CO2 is much more soluble than methane at the operational temperature employed in water scrubbing, thus reducing the risk of losing methane during scrubbing and consequently reducing methane slip. The combination of gas flow and water needed for scrubbing is a fundamental parameter to evaluate for the maximization of process efficiency, defined as a liquid/gas ratio. The flow of water strictly depends on gas pressure, and thus gas flow, and on process temperature, because a difference of a few degrees may significantly change the amount of water needed for CO2 absorption, as shown above. Table 3.2 shows a selection of liquid/gas ratios found in recent literature. A feature of the widely used water scrubbing technique is that it can simultaneously remove H2S from a biogas stream: hydrogen sulfide is even more water-soluble than CO2, and the same pressure and temperature parameters are effective for both contaminants (Deublein and Steinhauser, 2010). At the same time, the scrubbing of hydrogen sulfide means the water solution cannot be easily regenerated without boiling the solution at temperatures up to 300
C under vacuum (Deublein and Steinhauser, 2010), thus causing the production of significant amounts of wastewater and the need for the

TABLE 3.1 Henry’s constant for the main gases involved in the water scrubbing of biogas (Sander, 2015). Species name

Hcp at 298.15 K (mol/m23 Pa21)

d ln Hcp d ð1=T Þ

CO2

3.4 3 1022

2400

CH4

25

1600

23

2100

H2S

1.4 3 10

1.0 3 10

TABLE 3.2 Selected liquid/gas ratios for water scrubbing purposes. Liquid/gas ratio 0.140.5 (Xiao et al., 2014) 0.20.25 (Walozi et al., 2016) 1.02.3 (Vijay et al., 2006; Geng et al., 2015) 2.65 (Singh and Dwivedi, 2019)

(K)

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continuous addition of new, clean water. For this reason, H2S is usually removed before CO2, for example with activated carbon filters, by adsorption on iron oxides, by precipitation with iron salts or by bioscrubbing (Paolini et al., 2019; Munoz et al., 2015). A pressure of 1.2 MPa was reported as the one with the highest CO2 removal rate (94.2%) with a liquid/gas ratio of 0.5 (Xiao et al., 2014), while a pressure of 0.74 MPa and a liquid/gas ratio of 3.9 resulted in a methane purity of 95.1% (Singh and Dwivedi, 2019), confirming the interrelation of these two parameters. Another study confirmed this trend, reporting that a pressure of 1 MPa and a liquid/gas ratio of 1.2 resulted in a CO2 absorption of 99% (Vijay et al., 2006). Several studies (Geng et al., 2015; Alutu and Maduegbunam, 2017; Walozi et al., 2016) have explored the possibility of employing the water scrubbing technique at lower or ambient pressures in order to reduce costs, which could surely enhance the impact of biogas production and use in rural and developing areas. Once pressure is set at ambient values, other parameters can be manipulated to optimize the process, such as packing depth, column size, and liquid/gas rate. Purification results are predictably lower than those obtainable with high-pressure treatments, but an increase in methane content from 55% to 80% is nonetheless reported.

3.2.2

Absorption in NaOH solutions (alkaline scrubbing)

Absorption of CO2 can be significantly improved by using an alkaline solution instead of water. For CO2 absorption, in general, it is a well-known process (Kucka et al., 2002), while some research efforts are still needed for its application in biogas upgrading, as shown in this subsection. The chemical reactions involved when the hydroxide ion is provided by sodium hydroxide are: CO2 1 NaOH-NaHCO3 NaHCO3 1 NaOH-Na2 CO3 1 H2 O For the purification of biogas from CO2 via alkaline absorption, an absorption column is usually employed: these columns can be described as vertical cylinders filled with an alkaline solution and packed with small objects, usually metallic or plastic and of random shapes, in order to augment the gas/liquid interface area. Biogas flows from the lower section of the column and bubbles naturally upward, while the alkaline solution flows downward by gravity. Along this path to the top of the column, CO2 molecules interact with sodium hydroxide molecules as shown above (Tippayawong and Thanompongchart, 2010). Solution flow is critical in the operation of an absorption column, as too high a flow will result in flooding, while too low a flow will result in excessive discharge (Gabelman and Hwang, 1999).

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The chemical reactions involved in the absorption of CO2 depend on several parameters, such as concentration of both CO2 and alkaline species, solution pH, biogas pressure and the height of the column (Tippayawong and Thanompongchart, 2010; Malla et al., 2016). Theoretically, 1 ton of CO2 could be absorbed with 0.9 tons of NaOH (Yoo et al., 2013). One problem of this technology is the relative difficulty of regenerating the alkaline solution: Na2CO3 can be thermally decomposed to Na2O, which is a precursor to NaOH, but needs temperatures up to 800
C for this reaction to happen (Siriwardane et al., 2007). The regeneration of the alkaline solution can be facilitated by the addition of several compounds, such as CaO or Ca(OH)2, which decrease the needed temperature for the regeneration to happen (Yoo et al., 2013). On the other hand, NaOH is a relatively abundant and cheap compound, and it is also a waste material from other industrial processes, like chlorine gas production. Moreover, the resulting sodium carbonate could be reused in the ceramic, glass, or soap-making industries (Luis et al., 2013). The efficiency of the absorption of CO2 with NaOH solutions is generally high. A 0.1 M concentration of NaOH in water solution fed to a packed tower with a surface/volume ratio of about 2000 m2/m3 was able to upgrade biogas from a 46.8% concentration of CO2 to a posttreatment concentration of 3.2%, and saturating the solution in 100 min (Tippayawong and Thanompongchart, 2010). Another study showed a 15% CO2 removal efficiency by a 0.2 M NaOH solution, that is, at pH 13: the addition of an oxidant resulted in a slight decrease of pH and a near doubling of the efficiency (Dubois and Thomas, 2010).

3.2.3

Absorption in K2CO3 solutions (hot potassium carbonate)

In order to overcome the abovementioned technical barriers of alkaline scrubbing, hot potassium carbonate has been recently proposed for CO2 absorption from biogas. The hot potassium carbonate process is another chemical absorption process where potassium carbonate in solution is mixed with a gas stream containing CO2, which is absorbed as shown in the following reversible reaction: K2 CO3 1 CO2 1 H2 O-2KHCO3 Absorption occurs at high pressure (around 3000 kPa) and temperature (up to 120
C) (Smith et al., 2012), and consequently the resulting solution can be easily regenerated just by lowering the pressure or, to speed up the process, by also slightly raising the temperature, thus freeing the captured CO2 in a reaction that is the exact reverse of the capture reaction (Paolini et al., 2019; Li et al., 2014). High temperatures also raise the solubility of potassium carbonate in water, allowing the use of very concentrated solutions. Similarly to the other described processes, the hot potassium carbonate solution washes the biogas stream countercurrent in a packed column, where

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the biogas flows upwards and the liquid goes downwards by gravity (Smith et al., 2012; Zhao et al., 2011). In general, the same considerations already set out for the other processes also apply in this case, in particular regarding column packaging and the flow of gas and solution to avoid flushing or flooding. Likewise, models show the applicability of hollow fiber membrane contactors to hot potassium solutions (Mehdipour et al., 2013; Saidi, 2017). The use of potassium carbonate as absorber has several advantages over other alkaline compounds, first of all the ease of regenerating the solution (Kamps et al., 2007), which only requires moderate heating and lowering the pressure for captured CO2 to be let out as gas into the atmosphere. Moreover, at temperatures ranging from 40
C to 120
C the absorption capacity of potassium carbonate is already high, with a ratio up to 3.6 CO2/K2CO3 (Paolini et al., 2019). Another important advantage of this technology is the low rate of methane slip, which will be covered in detail in Section 3.6. Finally, potassium carbonate is inexpensive, less toxic than other solutions, much less toxic than amines, and also less corrosive and less prone to degradation in the presence of oxygen (Smith et al., 2012). Hot potassium carbonate is a well-known methodology for the purification of syngas and for the abatement of CO2 in off-gas. Up to 56% CO2 removal from syngas starting with a 12% CO2 content is reported with a gas flow rate of 300 kg/h, solvent flow rate of 1800 kg/h and 20% concentration, temperature of 100
C and pressure around 600 kPa (Smith et al., 2012). Another case, related to an integrated gasification combined cycle using coal, reports a CO2 removal efficiency of 37% at 150
C and 3000 kPa (Urech et al., 2014). Despite its strong potential for CO2 removal, there is still a knowledge gap for the application of this methodology to the upgrading of biogas to biomethane. Aside from simulation studies, only one experimental study is known at this moment, which reports a CO2 content in cleaned biomethane of 1.2% from a starting point of 39% in biogas (Paolini et al., 2019). This absorption reactor was working at 5001200 kPa and 60
C90
C, with a biogas flow rate of 150200 m3/h and solvent solution flow rate of 2.53 m3/h.

3.3 3.3.1

Absorption configurations Packed column reactors

The packed column absorption reactor is the most common configuration for biogas upgrading by water absorption. In its essential aspect, it is a vertical tank (tower) in which the gas and the liquid flow against each other: the liquid solution flows downwards while the gas rises upwards, each moved by the action of gravity. To maximize the contact between liquid and gas, the tower is generally filled with small objects that force the liquid to percolate

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slowly through tortuous paths and the gas to rise in small bubbles. Since the liquid and the gas compete for space, very intimate contact occurs between the two and there is a very large contact area through which the transfer of chemical species can take place. The countercurrent movement with the steam moving upward and the liquid downward is very practical for two reasons: it occurs naturally by gravity while the liquid flows downwards and the gas rises through it, and the gas meets a less and less charged liquid while becoming progressively cleaner as it heads upwards (Fig. 3.1). The exact structure of the scrubbing tower and the selection of the solvent strictly depend on the composition of the biogas that has to be cleaned. In the specific case of CO2 stripping performed in a packed column, like for any other chemical scrubbing of a gas stream, it depends on the selection of a solvent, of the correct packing, and of gas and liquid flow rates (Abdeen et al., 2016), and on the different solubilities of methane and carbon dioxide in water, with CO2 having a much higher solubility (Bauer et al., 2013).

3.3.2

Hollow fiber membrane contactors

Another reliable design for the alkaline scrubbing of CO2 from biogas is hollow fiber membrane contactors. Membrane contactors are hydrophobic, highly porous, thin films organized in concentric cylinders. The biogas stream flows in the fiber lumen, while the alkaline solution flows in

FIGURE 3.1 Schematic illustration of a water scrubber.

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FIGURE 3.2 Sample hollow fiber membrane contactor provided with a baffle (Gabelman and Hwang, 1999).

countercurrent in the shell layer. In this configuration, the gas/liquid interface forms only at the mouth of the pores, and the exchange of CO2 will depend only on concentration, instead of the pressure or flow rate (Gabelman and Hwang, 1999). Fouling of the membrane can be a problem in the case of particularly dirty biogas, with suspended particles, but this risk is mitigated by the lack of convective flow through the pores (Fig. 3.2). The main advantage of membrane contactors over packed columns is the greatly increased contact surface area between gas and liquid, up to 10 times higher (Chabanon et al., 2015), with consequent volume reduction of the apparatus and reduced energy needs. Hollow fiber membrane contactors also need baffles built into their design, to ensure turbulent flow, ensuring a more efficient mass transfer, while in packed columns turbulent flow occurs naturally.

3.4

Chemical promoters in water absorption

In each examined methodology, one of the reactions involves the chemical absorption of CO2 gas in a water media, as follows: CO2 1 OH2 -HCO2 3 Being a slow reaction, it is the limiting step of the process, whether the hydroxide ion is provided by water itself or by a strong base such as NaOH or KHCO3. Chemical promoters can significantly speed up this reaction by acting as self-regenerating intermediaries involved in turning carbon dioxide into carbonate. The process can be promoted by ammonia (Rahimpour and Kashkooli, 2004; Mahmoodi and Darvishi, 2017), inorganic salts such as borate (Smith et al., 2012;

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Ahmadi et al., 2008), and organic amines such as monoethanolamine (MEA) (Thee et al., 2014; Thee et al., 2012), diethanolamine (DEA) (Saidi, 2017; Saidi et al., 2014), and piperazines (Park, 2014; Kim et al., 2016). In this type of catalysis, the promoter first forms an intermediate with carbon dioxide, and the intermediate is then hydrolyzed to bicarbonate, which is the final step. Borate, for example, can promote the absorption of CO2 via the following reactions: 1 BðOHÞ3  H2 O2BðOHÞ2 4 1H 2 BðOHÞ2 4 1 CO2 -BðOHÞ4 CO2 2 BðOHÞ4 CO2 2 1 H2 O-BðOHÞ3  H2 O 1 HCO3

Boric salts are relatively environmentally safe and only small amounts are needed for significant effects, increasing carbon dioxide absorption up to 100% by adding 3%4% of borate in weight to the solution. CO2 absorption by alkaline solution is very pH-sensitive though, and the addition of boric acid to the solution can cause substantial pH drops, which might even nullify the effect of the borate addition in an inaccurate reactor design (Smith et al., 2012). Ammonia and its amine derivatives such as MEA, DEA, and piperazine act on the same rate-limiting reaction in a different way: CO2 1 RR0 NH-RR0 NCOOH 0 RR0 NCOOH 1 OH2 2HCO2 3 1 RR NH

The addition of amines as promoters to potassium carbonates combines the advantages of both chemical absorbers, that is the large absorption capacity of potassium carbonate with the high absorption rates of amines, which would otherwise be lower due to the low concentration of hydroxyl ions (Saidi, 2017). Amine concentration is a fundamental parameter, but the effect on carbon dioxide capture stops increasing over a threshold value, which was found to be around 10% wt./wt. (Rahimpour and Kashkooli, 2004). The addition of a 2.2 M solution of MEA, for example, can accelerate the absorption rate of a 30% potassium carbonate solution by a factor of 45
C at 63
C, and an increase in temperature can accelerate it further (Thee et al., 2012). Piperazine is generally used together with alcoholamines MEA and DEA to further accelerate CO2 absorption, since the rate constant of piperazine with carbon dioxide is one order of magnitude higher than that of MEA (Dang and Rochelle, 2003). Piperazine can also be used alone as the absorber, and a 10% piperazine solution can reach CO2 absorption levels 20 times higher than 30% DEA solutions (Dubois and Thomas, 2009). Piperazine has low solubility in water, limiting its usability as it cannot be highly concentrated; on the other hand, slight amounts of piperazine added to other amine solutions can

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significantly increase the absorption capacity and rate. The absorption rate of a piperazine-promoted MEA solution, for example, has a 2.5 times higher absorption rate than a similar unpromoted solution (Dang and Rochelle, 2003). The effects of a 5% piperazine solution on the absorption rate of potassium carbonate were reported equivalent to those of a 30% MEA solution, as the moles of CO2 absorbed per mole of solute are higher for piperazine (Kim et al., 2016). As for sarcosine, an amino acid with chemical properties similar to amines, a 2 M solution at 60
C added to a 30% K2CO3 solution accelerated the absorption of CO2 by a factor of 179 (Thee et al., 2014). Slightly lower effects are reported for glycine and proline, while a rise in temperature up to 82
C corresponds to an increase in CO2 absorption rate for these promoters, which are generally comparable with or more effective than MEA and DEA.

3.5

Energy consumption

Energy consumption, defined as the electricity and heat necessary for the upgrading of each cubic meter of biogas, varies with size for CO2 scrubbing plants. In general, electricity consumption values are given in Table 3.3. Electricity consumption falls from 0.3 to 0.23 kWh/m3 as the plant capacity increases, while there is a substantial difference between winter and summer depending on the need for cooling the water during the summer (Bauer et al., 2013). Gas and water compression, together with water cooling, can add an average of 0.2 kWh/m3 more to the energy needs of a plant (Munoz et al., 2015). As water scrubbing plants can be significantly different from each other per several parameters such as design, input gas purity, output gas desired purity, or reagents employed, it is difficult to make broad generalizations, as plant builders supply energy consumption values as high as 0.9 kWh/m3 (Sun et al., 2015). The energy efficiency of methane upgrading may be calculated with the following formula: output Energy efficiency ð%Þ 5 energy energy input 3 100 In this case, one standard cubic meter of upgraded methane has an energy content of 21 kWh/Sm3 (Speight, 2011). TABLE 3.3 Electricity consumption for water scrubbing. Electricity consumption (kWh/m3) 0.20.3 (Bauer et al., 2013; Singhal et al., 2017) 0.40.5 (Munoz et al., 2015; Wesley Awe et al., 2017) 0.50.9 (Sun et al., 2015) 0.20.43 (Andriani et al., 2014)

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Other costs associated with chemical compounds and the regeneration of scrubbing solutions vary, even if reagent costs are often negligible in comparison to other expenditures related to plant operation (Munoz et al., 2015).

3.6

Methane slip and efficiency

Methane slip represents the amount of methane that incorrectly flows out of upgrading plants together with off-gas due to limitation of the upgrading method or faulty equipment. Methane slip is a significant problem because of loss of profits and because of the impact of methane as a greenhouse gas 25 times more harmful than carbon dioxide. High methane slip rates also require further treatment of the off-gas before release into the atmosphere. Table 3.4 shows the average methane slip for water scrubbing technologies. Considering the different solubilities in water for carbon dioxide and methane, water scrubbing is generally characterized by methane losses below 2%, as reported by several biomethane upgrading plant suppliers (Sun et al., 2015). In addition, while methane slip rates as high as 18% are reported under regular operation, they are caused by nonoptimized plants and flash tanks (Munoz et al., 2015; Kvist and Aryal, 2019). While generally low in water scrubbing upgrading, methane slip should still be mitigated down the line for its negative impact on the environment. Solutions from other industry fields, such as incineration, naval transport, and chemical plants, can already be employed for biogas upgrading. Two of the most common solutions are afterburners, which are the cheapest and may be easily integrated, and regenerative thermal oxidizers, where heating and combustion of methane to carbon dioxide and water is coupled with the recovery of thermal energy (Kvist and Aryal, 2019). The efficiency of CO2 removal from the biogas stream, measured as the percentage of methane in the upgraded gas flow, is generally high, and reported as over 97% in most cases (Munoz et al., 2015; Singhal et al., 2017; Wesley Awe et al., 2017; Andriani et al., 2014). Efficiency also depends on

TABLE 3.4 Methane slip percentages for water scrubbing. Methane slip (%) 1 (Bauer et al., 2013; Singhal et al., 2017) ,2 (Munoz et al., 2015; Wesley Awe et al., 2017; Andriani et al., 2014; Kvist and Aryal, 2019) 12 (Sun et al., 2015)

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methane slip, as methane lost in off-gas translates into an effectively lower biomethane production.

3.7

Conclusions

The upgrading of biogas is an important technology when taking into consideration the need for more sustainable and environmentally friendly energy production means or fuels for vehicular transport. Several technologies already exist. This chapter focused on water scrubbing, which is a mature, well-tested, and widespread method for the removal of carbon dioxide from biogas capable of supplying high-quality biomethane suitable for injection in natural gas grids or for the automotive sector. Despite water scrubbing being one of the oldest biogas upgrading technologies, there are still areas where efficiency can be increased and costs lowered, such as the use of membrane contactors, hot potassium carbonate treatment, and chemical promoters. On the other hand, methane slip could quickly become the limiting factor in carbon dioxide removal technologies due to climate change and economic concerns; consequently, water scrubbing might become a leading biogas upgrading technology, due to its relatively lower methane slip.

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potassium carbonate and boric acid promoted potassium carbonate solvent. Int. J. Greenh. Gas Control 10, 6473. Available from: https://doi.org/10.1016/j.ijggc.2012.05.018. Speight, J.G., 2011. Combustion of hydrocarbons. Handbook of Industrial Hydrocarbon Processes. Elsevier. Strevett, K.A., Vieth, R.F., Grasso, D., 1995. Chemo-autotrophic biogas purification for methane enrichment: mechanism and kinetics. Chem. Eng. J. Biochem. Eng. J. 58, 7179. Available from: https://doi.org/10.1016/0923-0467(95)06095-2. Sun, Q., Li, H., Yan, J., Liu, L., Yu, Z., Yu, X., 2015. Selection of appropriate biogas upgrading technology—a review of biogas cleaning, upgrading and utilisation. Renew. Sustain. Energy Rev. 51, 521532. Available from: https://doi.org/10.1016/j.rser.2015.06.029. Sweeten, J.M., Fulhage, C., Humenik, F.J., 1981. Methane gas from swine manure. Leafl. Agric. Ext. Serv. 1860. Thee, H., Suryaputradinata, Y.A., Mumford, K.A., Smith, K.H., da Silva, G., Kentish, S.E., et al., 2012. A kinetic and process modeling study of CO2 capture with MEA-promoted potassium carbonate solutions. Chem. Eng. J. 210, 271279. Available from: https://doi.org/ 10.1016/j.cej.2012.08.092. Thee, H., Nicholas, N.J., Smith, K.H., da Silva, G., Kentish, S.E., Stevens, G.W., 2014. A kinetic study of CO2 capture with potassium carbonate solutions promoted with various amino acids: glycine, sarcosine and proline. Int. J. Greenh. Gas Control 20, 21202222. Available from: https://doi.org/10.1016/j.ijggc.2013.10.027. Tippayawong, N., Thanompongchart, P., 2010. Biogas quality upgrade by simultaneous removal of CO2 and H2S in a packed column reactor. Energy. 35 (12), 45314535. Available from: https://doi.org/10.1016/j.energy.2010.04.014. Urech, J., Tock, L., Harkin, T., Hoadley, A., Mare´chal, F., 2014. An assessment of different solvent-based capture technologies within an IGCC-CCS power plant. Energy. 64, 268276. Available from: https://doi.org/10.1016/j.energy.2013.10.081. Vijay V.K., Chandra R., Subbarao P.M.V., Kapdi S.S., 2006. Biogas purification and bottling into CNG cylinders: producing Bio-CNG from biomass for rural automotive applications. In: Second Joint International Confernce on Sustainable Energy and Environment. Available from: https://doi.org/10.1109/LPT.2009.2020494. Walozi, R., Nabuuma, B., Sebiti, A., 2016. Application of low pressure water scrubbing technique for increasing methane content in biogas. Univers. J. Agric. Res. 4 (2), 6065. Available from: https://doi.org/10.13189/ujar.2016.040206. Wesley Awe, O., Zhao, Y., Nzihou, A., Minh, D.P., Lyczko, N., 2017. A review of biogas utilisation, purification and upgrading technologies. Waste Biomass Valori. 8, 267283. Available from: https://doi.org/10.1007/s12649-016-9826-4. Wheatley, A.D., 1984. Biotechnology and effluent treatment. Biotechnol. Genet. Eng. Rev. 1, 261309. Available from: https://doi.org/10.1080/02648725.1984.10647788. Xiao, Y., Yuan, H., Pang, Y., Chen, S., Zhu, B., Zou, D., et al., 2014. CO2 removal from biogas bywater washing system. Chin. J. Chem. Eng. 22 (8), 950953. Available from: https://doi. org/10.1016/j.cjche.2014.06.001. Yoo, M., Han, S.J., Wee, J.H., 2013. Carbon dioxide capture capacity of sodium hydroxide aqueous solution. J. Environ. Manage. 114, 512519. Available from: https://doi.org/10.1016/j. jenvman.2012.10.061. Zhao, X., Smith, K.H., Simioni, M.A., Tao, W., Kentish, S.E., Fei, W., et al., 2011. Comparison of several packings for CO2 chemical absorption in a packed column. Int. J. Greenh. Gas Control 5 (5), 11631169. Available from: https://doi.org/10.1016/j.ijggc.2011.07.006.

Chapter 4

Factors affecting CO2 and CH4 separation during biogas upgrading in a water scrubbing process Rimika Kapoor, Pooja Ghosh and Virendra Kumar Vijay Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India

Chapter outline 4.1 Introduction 73 4.2 Approaches for CO2 removal from biogas 74 4.3 Water scrubbing technology 75 4.4 Water as a solvent for gases 76 4.5 Solubility of biogas components in water 77 4.6 Factors affecting biogas upgrading in water scrubbing process 78 4.6.1 Effects of operating parameters on CO2 removal in water scrubber 78 4.6.2 Effect of packed-bed design parameters 80

4.1

4.7 Scrubbing column internals 4.7.1 Packing support and gas distributor 4.7.2 Liquid distribution and redistribution 4.7.3 Demister or entrainment eliminator or mist eliminator 4.8 Major challenges and future directions 4.9 Conclusion Acknowledgments References

83 83 86

86 86 87 88 88

Introduction

The rising demand for energy and to curtail environmental impacts of fossil fuels has meant that energy systems fueled by sources that are more efficient, cost-effective, and reduce environmental emissions are in major demand. This search has led to biogas, a significant fuel among the various biomassderived fuel options. It is a clean and renewable gas obtained from anaerobic fermentation of organic biomass waste (Pertl et al., 2010). Biogas is the product of anaerobic digestion of organic waste. It mainly comprises of CO2 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00004-0 © 2021 Elsevier Inc. All rights reserved.

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(35% 45%) and CH4 (55% 65%), with traces of H2S, water vapor, and other trace compounds. Worldwide biogas is best known for being extensively used for heating and/or electricity generation (Bauer et al., 2013a,b). To increase the applicability of biogas to high-grade applications like transport and grid injection (injection of biomethane into the natural gas grid) as a substitute to fossil-based natural gas, its calorific value has to be increased by varying its composition. The process of removal of CO2 from biogas, also called upgradation or biomethane production, thus plays an important role not only in widening the scope of utilization of biogas to higher value applications same as natural gas but also in reducing the dependence on conventional fossil fuels (Persson et al., 2006).

4.2

Approaches for CO2 removal from biogas

Biogas upgradation is basically a gas separation process, yielding a CH4-rich gas commonly known as biomethane or bio-CNG. A variety of methods are available for separating CO2 and other unwanted constituents from biogas. However, choosing the appropriate method depends upon various factors like cost, energy requirement, site conditions, application, etc. (Wellinger et al., 2013). These methods are based on the differences in their molecular properties or the thermodynamic and transport properties of the components in biogas such as absorption (pressurized water scrubbing, physical or chemical absorption), adsorption (pressure swing adsorption), membrane (high pressure, low pressure) separation based on temperature, and pressure cryogenic upgrading (Andriani et al., 2014; Ryckebosch et al., 2011). Absorption can be either a physical or chemical process in which CO2 requires a solvent to absorb. Physical absorption employs solvents like water, polyethylene glycol, Selexol, and Genosorb based on the solubility of CO2 in the solvent, whereas the chemical absorption method uses chemical solvents such as mono-ethanol amine or di-methyl ethanol amine, which are highly CO2 selective and react chemically to drive into the solution (Tippayawong and Thanompongchart, 2010). The major constraints in regard to chemical solvents are toxicity, corrosivity, and significant energy consumption requirement for regeneration (Lin and Shyu, 1999). In the adsorption method, pressure swing adsorption (PSA), CO2 is adsorbed onto a porous solid surface such as activated carbon and then desorbed by changes in pressure. This method involves high methane losses, complicated process design, and high operation costs (Cavenati et al., 2005). Membrane separation depends upon selective permeability and the preferential transfer of CO2 from biogas through a semipermeable membrane, acting as a barrier, whereas other components are retained. However, the high pressures during the process, chemical corrosion and contamination of the membrane, and higher maintenance costs are some of its major drawbacks in addition to the high costs of the membrane (Shao et al., 2012). Cryogenic separation separates CO2 from biogas by cooling and compressing the gas. This process relies on the fact that

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different impurities liquefy in different temperature and pressure conditions (Goffeng, 2013; Jonsson and Westman, 2011). Of all the upgrading technologies, the water scrubbing technique based on the physical absorption process remains the simplest, and the most efficient, sustainable, and implemented method for biogas upgrading.

4.3

Water scrubbing technology

The water scrubbing method is a physical absorption process, that is, absorption without a chemical reaction. It is an upgrading process involving separation of gases based on the greater water solubility of CO2 in comparison to CH4 (Patterson et al., 2011). This process is governed by the principle of solubility of gases in water, which depends upon pressure, temperature, and the concentration of the gas in water. The water scrubbing process depends upon not only the properties of water as a solvent and solubility of gases in water but also on the various operative and packed-bed design parameters such as gas and water flow rates, pressure inside the column, packing material, diameter, and height of the packed-bed column. This lines up a crucial deliberation for the optimum design of a high-pressure water scrubbing-based packed-bed absorption process for CO2 removal biogas upgradation. The water scrubbing process employs a packed-bed column to provide a high surface area to encourage high mass transfer efficiency along with a low-pressure drop for CO2 absorption from biogas (Cozma et al., 2013a; Kohl and Nielsen, 1997; Billet, 1995). During the upgradation process, compressed biogas is fed from the bottom of the scrubbing column and water is fed from the top, as shown in Fig. 4.1. A counter current flow is facilitated between the water and biogas. As biogas flows up into the column, CO2 along with some CH4 is absorbed in water which exits the scrubbing column from the bottom. Water can be regenerated after depressurizing at low pressure and then recirculated back to the absorption column. This method involves lower capital investment and operational costs, is reliable and easy to maintain, as well as producing high gas purity and yield (Cozma et al., 2013b). Water scrubbing is a well-developed technology and is being increasingly implemented around the world. More than 500 biogas upgrading plants were in operation until the end of 2019, with over 200 plants based on water scrubbing technology (IEA Bioenergy, 2013), however, reports and literature reporting or investigating the effects of various operating and design factors on the performance of water scrubbing technology to achieve maximum CO2 removal from biogas are rare. The primary objective of this study is to succinctly review the effects of various parameters affecting the water scrubbing process to enhance CO2 removal efficiency. The study aims at promulgating essential information, research directions, and guiding principles for the water scrubbing process to upgrade biogas.

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FIGURE 4.1 Schematic diagram of a water scrubbing column.

4.4

Water as a solvent for gases

Despite the fact that solubility of CO2 is higher in chemical and physical solvents, water remains a preferred candidate for the separation of CO2 from biogas using scrubbing columns.

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Water as a solvent is innocuous, nonflammable, and chemically stable, with a low freezing point. It is relatively nonvolatile, noncorrosive, nonfoaming, and nonviscous, which increases the rate of absorption and provides low-pressure drop characteristics (Kestin et al., 1978; Houghton, 2008). Water is the most abundant and generally termed as a universal solvent with high solubility for polar gases and low solubility for nonpolar gases. The heat capacity of water is anomalously high with low viscosity at high pressures and low temperatures. The principal benefits of water as an absorbent for gases like CO2 are its solubility, abundant availability, and low cost. In addition, regeneration of pressurized water does not require energy or specialized equipment. This factor is enough to make water worth considering over chemical and organic physical solvents.

4.5

Solubility of biogas components in water

The solubility parameters of a gas are temperature and pressure. Solubility also depends upon polarity and van der Waals’ volume or chemical bonds of molecules. According to Namiot (1967), a molecule of CO2 is larger (has a higher molecular weight), more polarizable, and creates stronger intermolecular forces with water than CH4, thereby increasing its solubility in water. Precise knowledge of the solubility of the biogas constituents (CO2, CH4, and H2S) is important for designing and optimizing a water scrubbing process. Extensive literature is available on the solubility of gases in water but for the water scrubbing biogas upgradation process the results under low and medium temperatures and pressures (temperatures between 0 C and 60 C and pressures up to 50 bar) have to be considered. A wide range of solubility data are available for CO2, H2S, and CH4, as single-gas solubilities in water, aqueous solutions, electrolyte solutions, and organic solvents (Clever and Young, 1987; Duan et al., 2007). However, only a few studies have been published on the cosolubility of CO2 and CH4 in water. Cozma et al. (2013a) studied several thermodynamic models to calculate the solubility of CO2, CH4, H2S, N2, and O2 in water and found that they were consistent with the experimental data as reported by Huang et al. (1985). The solubility of CO2 in water is one of the most studied phenomena in all of physical chemistry. Calix et al. (2008) stated in their study that pressure has a direct and temperature an indirect effect on the solubility of CO2 in water (pressure increases the solubility of CO2). Methane solubility measurements in pure water are very low, and accurate measurements of the CH4 water system over a wide range of temperatures and pressures are sporadic. H2S, produced typically by anaerobic digestion of organic matter in sewers, is the one of the gaseous components of biogas. Due to its corrosive nature, removal of H2S from biogas is essential. In the water scrubbing process, selective removal of H2S along with CO2 takes place, since it is more soluble than CO2 in water (Carroll and Mather, 1989). At 0 C, CO2 has a molar concentration approximately 29 times greater than CH4, although this ratio reduces with temperature to approximately

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TABLE 4.1 Comparison of the solubility and Henry’s constant of biogas components at 25 C and 1 atm (Perry and Green, 1997; Sander, 2011; Uttamote, 2011). Gases

Solubility (g/L)

Henry’s law constant (M/atm) 22

Mole fraction

H2S

3.43

8.7 3 10

1.65 3 1023

CO2

1.45

3.6 3 1022

0.609 3 1023

CH4

0.0227

1.5 3 1023

0.277 3 1024

23:1 at 30 C (Sloan and Koh, 2008). At standard temperature and pressure (STP), the solubility for H2S is approximately 2.5 and 73 times higher than CO2 and CH4, respectively, as shown in Table 4.1. The water scrubbing process takes advantage of the higher water solubility of H2S and CO2 compared to CH4, thereby separating these gases with high efficiency.

4.6 Factors affecting biogas upgrading in water scrubbing process Removal of CO2 from biogas in a water scrubbing column is not only dependent on the dissolution properties of water as a solvent and solubility of gases in water but also on various operating and design parameters. Operational problems like a high-pressure drop or flooding phenomena are usually experienced during the process of gas separation in packed-bed absorption columns which hamper effective CO2 separation (Sen et al., 2013). Therefore, to avoid such problems to achieve high CO2 removal efficiency from biogas, there is a need to have in-depth knowledge of the packed-bed absorption column design and operating conditions. An insight into the various factors is discussed in the following subsections.

4.6.1 Effects of operating parameters on CO2 removal in water scrubber 4.6.1.1 Pressure Pressure plays a very crucial role in absorption of CO2 in water. La¨ntela¨ et al. (2012) observed an increase in CO2 removal efficiency from 85.8% to 88.9% as the pressure increased from 20 to 25 bar at a constant temperature of 10 C 15 C. Xiao et al. (2014) operated the packed-bed column at 8, 10, and 12 bar pressure and observed the highest CO2 removal efficiency of 94.2% at 12 bar pressure with only 2.6% of CO2 in upgraded gas. The solubility of CO2 increases in water as pressure increases, which results in a high

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CH4 concentration in biomethane. High pressure also reduces the quantity of water required for absorption.

4.6.1.2 Temperature Temperature is one of the most significant factors in the water scrubbing process. CO2 absorption in water is a physical process in which a large amount of dissolution heat is released (Carroll et al., 1991; Weiss, 1974). Low temperatures in the absorption column are thus advantageous for the enhancement of CO2 removal. An increase in the solubility of CO2 is observed at low temperatures. La¨ntela¨ et al. (2012) observed a decrease in CO2 removal efficiency from 88.9% to 87.3% as the temperature increased from 10 C 15 C to 20 C 25 C at a constant of 25 bar. A similar pattern was observed by Xiao et al. (2014). A decrease in CO2 removal efficiency from 85.3% to 52.2% was seen as the temperature increased from 7 C to 40 C. Therefore temperature control in the water scrubbing column for biogas upgradation should receive great attention, especially when the plant is operated during summer. 4.6.1.3 Water flow rate In a packed-bed absorption process, the water flow rate directly affects the CO2 removal efficiency. For the process of CO2 absorption in water the resistance to mass transfer lies in the liquid phase. High water flow rates provide a large amount of water molecules to react with CO2 molecules per unit time. There is an increase in the wetted surface of the packing material, which increases the effective area for mass transfer and thus reduces the liquid-film resistance (Richardson et al., 2002; Nozic, 2006). A high liquid flow rate also results in an increase in the absorption of gas because of the constant regeneration of the gas liquid interface (Ndiritu et al., 2013). Aroonwilas et al. (2001) credited enhanced absorption capacity to the increase in the concentration of water molecules reacting with carbon dioxide molecules causing formation of carbonic acid and greater wetting of the random packing. The solubility of the gas in water increases with pressure. This reduces the quantity of solvent required. Cozma et al. (2014) observed that higher pressure in the column compensated for a lower water flow rate as, at 17 bar, Vijay (2007) observed that the CO2 removal efficiency increased as the water flow rate increased at a constant pressure, and it was maximum at 1.8 m3/h water flow rate; above this flow rate, flooding started. This is because of the increase in gas pressure drop due to the reduced crosssectional area available for flow of gas resulting from the presence of increased water in the column (Treybol, 1968). 4.6.1.4 Gas flow rate The gas flow rate plays a significant role in sizing the packed-bed column as it directly affects the CO2 removal efficiency. At a very low gas flow rate in

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the column, the driving force inside the column is not high enough to cause efficient mass transfer (Ndiritu et al., 2013). As the gas flow rate increases, it causes better spreading capacity and enhanced mixing of the gas and water within the tower, causing mass transfer to increase. The driving force between CO2 and water is strengthened with increasing gas flow rates (Aroonwilas et al., 2001). The increase in the gas flow rate permits more CO2 molecules to transfer from the bulk gas phase to the gas liquid boundary layer, causing an increase in the mass-transfer coefficient (Tan et al., 2012). At much higher gas flow rates, gas pressure drop increases rapidly and water starts accumulating inside the tower (McCabe et al., 1993). Water accumulates in the upper section, almost preventing the flow of gas outside of the column and resulting in flooding (Sherwood et al., 1938). However, flooding can even occur at low gas flow rates but with very high liquid flow rates (Takahashi and Fujita, 1967). Hence, increased CO2 removal efficiency can be achieved by controlling pressure, and liquid and gas flow rates. Xiao et al. (2014) suggested in their study to apply relatively higher pressures and high liquid/gas ratios to achieve efficient CO2 removal. Chandra et al. (2012) observed a similar behavior with an increase in CO2 removal efficiency from 88.8% to 92.3% as the gas flow rate increased from 10 to 25 m3/h at a constant water flow rate of 1.5 m3/h and column pressure of 8 bar. With a further increase in the gas flow rate, CO2 removal efficiency decreased to 85.2% at a 25 m3/h gas flow rate due to increased pressure drop of the gas inside the tower.

4.6.2

Effect of packed-bed design parameters

4.6.2.1 Packing Packing in gas absorption processes plays a crucial role in providing a high surface area for mass transfer between gas and liquid. Packing not only provides resistance to the flow of gas and liquid inside the column but also increases the contact time between gas and liquid during the gas absorption process. Packing material should be chemically inert, robust, light weight, have a high void volume and surface area, and be cheap, mechanically, thermally and chemically stable, and provide a very low pressure drop. Packing material usually varies in size and shape and can be constructed of metal, graphite, plastic, clay, porcelain, or ceramic. Two main types of packing are used for gas absorption to meet the separation requirements of the specific application or process. Packing can be dumped or random packing (Pall ring, IMTP, Raschig rings) and structured or stacked packing (Flexipac, Mellapak, Gempak, Sulzer-BX) (Arachchige and Melaaen, 2012). Random packings are the discrete pieces of packings of a specific geometrical shape, which are basically dumped into the column. These have internal fingers, curves, and blades to boost interfacial gas liquid contact with

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negligible drag or hold-up. These are available in a size range of 6 75 mm (1/4 to 3 in.). Packings smaller than 25 mm are used mostly in laboratory or pilot-plant columns. Structured packing is formed from vertical sheets of ridged thin gage and an open honeycomb structure is formed with inclined flow passages. Structured packings are less commonly used than dumped packings and range from 50 to 200 mm (2 8 in.) (McCabe et al., 1993). Structured packings are predominantly designed for complex separations. The large contact area of structured packings provides a high overall mass transfer coefficient as compared to random packing (Fernandes, 2011). The dimensions of the scrubbing column are dependent upon the material and size of packing. A larger size of packing decreases the cost per unit volume and pressure drop per unit height of packed bed and thus the masstransfer efficiency. This results in a requirement for taller columns to achieve the required purity and separation (Treybol, 1968). For random packing it is recommended to use packing with a diameter 0.07 times that of the diameter of the packed-bed column. The liquid distribution and mass-transfer efficiency depreciates rapidly above this size of packing. Irregular distribution of packing inside the packed-bed columns causes an undesired process called channeling. This phenomenon takes place when the liquid moving down the column flows toward the region of greatest void space, which is generally near the wall where the packing is not tightly packed. Liquid flowing from the top toward the bottom of the packed-bed column has a normal propensity to flow toward the wall and the gas which enters from the bottom has a tendency to flow from the middle. These two conditions limit the contact between the liquid and the gas, thereby decreasing the mass transfer and pressure drop inside the column. To avoid channeling problems, flow rates of liquid and gas should be maintained according to the operating conditions and the number of packing elements per unit volume (packing density) should be distributed evenly in the column. Channeling is more severe in towers filled with structured packing, which is the main reason they are not very commonly used (Richardson et al., 2002). For CO2 absorption in water in packed-bed columns, random packings mostly are used. Kapdi et al. (2005) reported a randomly packed-bed column with locally available packing materials capable of removing 30% more CO2 by volume as compared to the scrubbing systems without a packed bed. Chandra et al. (2012) used randomly filled ceramic Resching rings in a packed-bed column and obtained 95% CO2 removal efficiency. Pall rings were used by Rasi et al. (2008) and La¨ntela¨ et al. (2012) in the packed-bed column for high CO2 removal efficiency from biogas.

4.6.2.2 Diameter The diameter describes and limits the capacity of the scrubbing column in terms of the gas flow rate. The diameter of the scrubbing column is

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TABLE 4.2 Suggested column diameter and packing sizes at different gas flow rates. Gas flow rate (N m3/h)

Column diameter (m)

Recommended size of packing (mm)

,50

,0.3

,25

50 500

0.3 0.9

25 38

.500

.0.9

50 75

calculated by its cross-sectional area. Normally, the column diameter is designed to ensure good liquid and gas distribution inside the column and to operate at the highest economical pressure drop. Normally, scrubbing columns are selected so as to process at a fraction of the flooding velocity. It is the velocity of the gas at which liquid droplets are carried along with the exiting gas stream (flooding condition). Usually, the columns are operated at 50% 75% of the flooding velocity, but the recommended design values for absorbers are 15 50 mm water per m (meter) of packing (Nozic, 2006). In practice, the packing diameter will not only depend on the physical properties and flow rates of the liquid and gas, but also on the uniformity of the liquid distribution throughout the column, which is dependent on the column diameter and height. Channeling is less prominent when the diameter of a single packing piece is smaller than at least 1/8 the tower diameter. Above this size, liquid distribution becomes inappropriate and masstransfer efficiency decreases. Generalized pressure-drop correlations are used to calculate the column cross-sectional area at a particular pressure drop per unit height of packing (McCabe et al., 1993). Table 4.2 presents the suggested size of the packing and the flow rate of gas for the selected diameter of the packed-bed column. Packings with small sizes are comparatively more expensive than the larger ones. Generally, the largest size of random packing recommended for large columns is 50 mm (Richardson et al., 2002; Treybol, 1968).

4.6.2.3 Height The height of the scrubbing column depends upon the level of purity required for gas separation (Keskinen et al., 1990). Height is dependent upon the number of transfer units (NTU) and height of a transfer unit (HTU). NTU is the degree of difficulty of separation (a larger NTU will be required to achieve very high product purity) and HTU is the measure of effective separation for a particular packing (inversely proportional to mass-transfer coefficient). The larger the mass-transfer coefficient, the smaller the value of HTU. A higher bed can separate more gas from the incoming gas mixture

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with lower incoming gas flow rates, whereas a wider column can treat a larger volume of biogas. According to Hunter and Oyama (2000), a height-to-diameter ratio of 10:1 20:1 for water flow rates of 0.04 2 m3/min/m2 of the column area is used for water scrubbing plants. Tynell et al. (2007) reported a packed-bed height of 10.0 m for a water scrubbing process operating at 9 12 bar pressure. Rasi et al. (2008) used instead of the conventional diameter/height ratio, a small height-to-diameter ratio of 3:1 (height 1.85 m, diameter 0.6 m) for low water flow rates of 5 and 10 L/min (0.3 0.6 m3/h) and high pressures of 10 30 bar to produce upgraded biogas with over 90% methane content. The higher pressures and low water flow rates compensated for the lack of column height. La¨ntela¨ et al. (2012) also used a height-to-width ratio of 3:1 for water flow rates of 5 L/min (0.3 m3/h) and 11 L/min (0.66 m3/h). The small packed-bed height was compensated with high pressures of 25 bar in the column with water recycling and observed the highest CO2 removal efficiency of 88.9%. The reason for low CO2 removal efficiency was attributed to low height-to-width ratio as compared to the conventional plants. Chandra et al. (2012) used a height of 3.5 mm and diameter of 0.15 m of the packedbed column. The height-to-diameter ratio of 20:1 was used to obtain a CO2 removal efficiency of 93.4% at 10 bar pressure (Table 4.3).

4.7

Scrubbing column internals

The proper selection scrubbing column internals such as packed-bed supports, liquid distributors and collectors, and other column internals is essential for optimum packing performance and efficient mass transfer. However, no experimental data are available on the selection of scrubbing column internals to enhance mass transfer.

4.7.1

Packing support and gas distributor

Support plates are provided to physically support the cumulative weight of the packing and the liquid held up during operation. These are designed to provide maximum open area for uniform gas distribution and continuous flow of the liquid through the plate for minimal pressure drop in the column. If a gas distributor is not selected properly it can create uneven gas liquid contact, channeling the gas through the column and causing build-up of small liquid pockets on the support plate. This decreases the column absorption efficiency. To create minimum pressure drop, a support plate with an open area greater than 70%, and preferably above 85%, should be selected. The factors affecting selection of the support plate are the column diameter, packing size and type, height of the packed bed, and the flow rate of the gas and liquid. The most commonly used packing support is the gas injection grid, with an effective area of 80% to more than 100% of the column cross-section.

TABLE 4.3 Summary of the effects of operating and design factors on CO2 separation from a gas mixture in a water scrubbing column. Factor

Results/conclusion

Pressure

G

G

Temperature

G G

Water flow rate

G G

G

Gas flow rate

G

G

G

G

G

Ref

Solubility of gas in water increases at high pressures which increases the rate of absorption of gas and thus decreases the quantity of solvent required. CO2 separation efficiency (CO2 absorption in water) increases with an increase in pressure.

Clever and Young (1987)

The dissolvability of CO2 decreases with an increase in temperature. Lower temperature in the absorption tower is beneficial for the enhancement of CO2 separation.

Kestin et al. (1978), Clever and Young (1987), Calix et al. (2008)

Low water flow rates cause less area for mass transfer of gas molecules. High liquid flow rate causes an increase in absorption of gas due to the continuous renewal of the gas liquid boundary layer causing an increase in the concentration of water molecules reacting with carbon dioxide molecules. At very high water flow rates gas pressure drop increases due to reduced crosssectional area available for flow of gas resulting from the presence of increased water in the tower.

Wilhelm et al. (1977), Lide (1993)

Important parameter for sizing the packed-bed column for maximum absorption of CO2 in water. When the gas flow rates in the column are too low, the driving force inside the tower is not high enough to cause efficient mass transfer. As the gas flow rates are increased, they cause better spreading capacity and enhanced mixing of the gas and water causing mass transfer to increase. At very high gas flow rates, gas pressure drop increases rapidly and water starts accumulating in the upper part of the column preventing the flow of gas and this results in flooding condition. Too high gas flow rates can cause flooding in the column, however with high liquid flow rates flooding can occur even at lower gas flow rates.

Clever and Young (1987), Calix et al. (2008), Namiot (1961)

Packing

G

G

G

Column diameter

G

G

Column height

G

G

The size of packing used influences the height and diameter of a column, the pressure drop, and cost of packing. Packings with small sizes provide larger area per unit packed volume which gives more pressure drop per unit bed height and hence increased mass transfer. As the packing size is increased, the cost per unit volume of packing and the pressure drop per unit height of packing are reduced, which reduces the masstransfer efficiency. This results in a requirement for a taller column height to achieve the required purity and separation.

Wen and Hung (1970), Rettich et al. (1981)

The diameter of the packed-bed column limits its capacity and is determined by its cross-sectional area. To avoid channeling in the column packed with random packings, the diameter of the individual packing pieces should be smaller than at least one-eighth the tower diameter. Above this size, liquid distribution, and hence the mass-transfer efficiency, depreciates rapidly.

Wilhelm et al. (1977), Morrison and Billett (1952)

The height of the packed-bed column is usually determined by the effectiveness of gas separation and the level of purity of gas required. A height-to-diameter of 10:1 20:1 for water flow rates of 0.04 2 m3/min/m2 of the column area is used for water scrubbing plants.

Rettich et al. (1981), Carroll and Mather (1997)

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4.7.2

Liquid distribution and redistribution

The liquid flowing downwards in the column tends to flow along the wall resulting in channeling. Such nonhomogeneities are unfavorable for high mass-transfer efficiency. For that reason, to avoid channeling, liquid distributors and redistributors are provided in the scrubbing column for an evenly wetted packing, good gas liquid contact, and hence satisfactory performance of the absorption system. The selection of a liquid distributor depends upon (1) flow rate of liquid, (2) fouling or plugging tendency, and (3) the turndown ratio. Long towers have a liquid redistributor to collect and direct the liquid off the column wall toward the center of the column (Pertl et al., 2010). Liquid redistributors are required at every 8 20 ft depth of random packing. Pan-, rick-, trough-, and V-weir-type distributors and redistributors are commonly used in scrubbing columns. Pan-type distributors are most commonly used for columns with small diameters. Generally, such problems occur at low flow rates of liquid if the liquid is not distributed evenly by the distributor on the packed bed. Also, a redistributor is located at a certain height in the packed bed which redistributes the liquid and vapor uniformly over the surface of the packed bed. Thus liquid redistributors are used to redirect the fluid flow toward the column center.

4.7.3

Demister or entrainment eliminator or mist eliminator

The gas exiting through the top of the column at high velocities may carry droplets of liquid as mist. To avert this, a mist eliminator can be installed at the top of the column in the form of corrugated sheets or a layer of mesh to collect the liquid droplets, which conjoin and fall back into the column. Sometimes, a layer of packing is also put above the liquid distributor. This layer, named “dry packing,” acts as an entrainment eliminator.

4.8

Major challenges and future directions

An interpretative literature review reveals that there is a strong possibility to improve the process of water scrubbing-based biogas upgradation systems. Being the most commonly implemented technology, thorough research and development for optimization of the process and investigation of the effects of various factors on the performance of the technology are urgently needed. An important point which needs attention is to keep various parameters within the desired range to increase the CO2 removal efficiency of the water scrubbing process as well as to practically maintain and monitor them thoroughly to reduce the energy consumption and operational costs of the upgrading plant. A major consequence of using water as a solvent for biogas upgradation is saturation of upgraded biogas with water vapor. To meet the standard for

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use of biomethane as a vehicle fuel or injection in the grid, water vapor should be within permissible limits. There are modest techniques available for removal of moisture from upgraded biogas which add to the cost and energy requirement of the process. Plugging or growth on packings in absorption columns is an existing problem in plants without regeneration of water. The problem also exists in plants with regeneration but not to the same extent. Clogging in the scrubber in the long run is another problem observed under field conditions. When the packings are plugged, they do not scatter the water, which causes difficulties in reaching a sufficient methane concentration. Plugging makes it more difficult for water to flow down in the column. There is an urgent need to explore novel packing materials and column internals which can enhance mass-transfer performance and comparatively lower pressure drop in the column. During the water scrubbing process, a small quantity of CH4 (1% 10% of biogas) is also lost in water as CH4 loss. As CH4 is a strong greenhouse gas and its loss is also negative for plant economics, losses should be kept as low as possible. To reduce CH4 loss into the environment, a few treatment technologies are available for the off-gas emanating from a biogas upgrading plant such as to use the off-gas as a lean fuel for combustion, and it also can be oxidized by thermal or catalytic oxidation, but these methods are unfeasible as they are complex, costly, require sophisticated design and careful sizing of the system, as well as another energy source for combustion. Furthermore, these techniques oxidize CH4 and do not recover CH4. Currently, one of the most preferred methods of CH4 loss recovery is using a flash tank to partly depressurize water at 2 4 bar to release off-gases. To achieve high biomethane quality as well as CH4 loss recovery with reduced energy cost, there is a need for improved design of scrubbing columns combined with a flash tank and optimization of operating conditions. The cost of biogas upgradation is influenced by scale, which has a direct impact on the economics. Water scrubbing-based biogas upgrading at small scale can be expensive. There is an urgent need to develop cost-efficient water scrubbing technology for small-scale applications. A cost-effective and energy-efficient process is achieved at higher scales of production. Many studies still need to be done to confirm the efficiency of the process when it is scaled up.

4.9

Conclusion

Of these technologies, water scrubbing is found to be the most feasible and cost-effective. It is a well-developed, easy, and simple technology, and is being increasingly acknowledged and implemented around the world. In spite of this, the related literature provides very little information concerning the influence of operating and design factors affecting the performance of water scrubbing technology to achieve maximum CO2 removal from biogas. Various operating and designing factors affecting CO2 removal efficiency have been discussed in this

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chapter. A thorough knowledge of the factors affecting efficient removal of CO2 from biogas and production of biomethane can be helpful in designing such simulation software to ease the tedious and complex theoretical procedure of designing packed-bed columns. Optimization of process and design parameters of water scrubbing columns for high CH4 recovery, reduction of capital and operating costs and energy consumption of biogas upgrading plants is an active area of research. Operational process parameters like pressure, flow rate, and temperature of the gas should be optimized to reduce the large quantities of water required, cost of biogas compression, and water pumping. Efficient methods of water regeneration and recycling are also areas worth exploring.

Acknowledgments We would like to express our sincere thanks to the Indian Institute of Technology Delhi for providing institute Post-Doctoral Fellowship to R. Kapoor and the Department of Science and Technology, Govt. of India for providing INSPIRE Faculty grant to P. Ghosh [DST/INSPIRE/04/2016/000362].

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Morrison, T.J., Billett, F., 1952. The salting-out of non-electrolytes. Part II. The effect of variation in non-electrolyte. J. Chem. Soc. 3819 3822. Namiot, A.Y., 1961. In: Clever, H.L., Young, C.L. (Eds.), Solubility Data Series, Methane. Pergamon Press, Oxford, pp. 27 28. Namiot, A.Y., 1967. Reasons for small solubility of nonpolar gases in water. J. Struct. Chem. 8, 363 366. Ndiritu, H.M., Kibicho, K., Gathitu, B.B., 2013. Influence of flow parameters on capture of carbon dioxide gas by a wet scrubber. J. Power Technol. 95, 9 15. Nozic, M., 2006. Removal of carbon dioxide from biogas. , http://www.chemeng.lth.se/exjobb/ E264.pdf . . (Accessed January 2015). Patterson, T., Esteves, S., Dinsdale, R., Guwy, A., 2011. An Evaluation of the policy and techno-economic factors affecting the potential for biogas upgrading for transport fuel use in the UK. Energy Policy 39, 1806 1816. Perry, R.H., Green, D., 1997. Perry’s Chemical Engineers’ Handbook. McGraw-Hill, New York. 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. , http://www.iea-biogas.net/files/daten-redaktion/download/publi-task37/ upgrading_report_final.pdf . . Pertl, A., Mostbauer, P., Obersteiner, G., 2010. Climate balance of biogas upgrading systems. Waste Manage. 30, 92 99. Rasi, S., Lantela, J., Veijanen, A., Rintala, J., 2008. Landfill gas upgrading with counter current water wash. Waste Manage. 28, 1528 1534. Rettich, T.R., Handa, Y.P., Battino, R., Wilhelm, E., 1981. 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. 85, 3230 3237. Richardson, J.F., Harker, J.H., Backhurst, J.R., 2002. Coulson and Richardson’s Chemical Engineering, fifth ed. Pergamon Press, Oxford. Ryckebosch, E., Drouillon, M., Veruaeren, H., 2011. Techniques for transformation of biogas to biomethane. Biomass Bioenergy 35, 1633 1645. Sander, R., 2011. Carbon dioxide. Henry’s Law data. NIST Chemistry Webbook. National Institute of Standards and Technology. Sen, D., Sarkar, S., Bhattacharjee, S., Bandopadhya, S., Ghosh, S., Bhattacharjee, C., 2013. 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. 52, 1731 1742. Shao, P., Dal-Cin, M., Kumar, A., Li, H., Singh, D.P., 2012. Design and economics of a hybrid membrane temperature swing adsorption process for upgrading biogas. J. Membr. Sci. 413 414, 17 18. Sherwood, T.K., Shipley, G.H., Holloway, F.A.L., 1938. Flooding velocities in packed columns. Ind. Eng. Chem. 30, 765 769. Sloan, D.E., Koh, C., 2008. Clathrate Hydrates of Natural Gases, third ed. Taylor and Francis, CRC Press, pp. 120 121. Takahashi, T., Fujita, K.A., 1967. Correlation of flooding velocities in counter current gas-liquid contactor of column type. Mem. Sch. Eng. Okayama Univ. 2, 43 49. Tan, L.S., Shariff, A.M., Lau, K.K., Bustam, M.A., 2012. Factors affecting CO2 absorption efficiency in packed column: a review. J. Ind. Eng. Chem. 18, 1874 1883. Tippayawong, N., Thanompongchart, P., 2010. Biogas quality upgrade by simultaneous removal of CO2 in a packed column reactor. Energy 35, 4531 4535.

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Treybol, R.E., 1968. Mass transfer operations, third ed. McGraw-Hill, New York. Tynell, A., Borjesson, G., Persson, M., 2007. Microbial growth on pall-rings a problem when upgrading biogas with the technique absorption with water wash. Appl. Biochem. Biotechnol. 141, 299 319. 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. , http://digi.library.tu.ac.th/thesis/siit/0258/title-appendix.pdf . . 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. , http://www.esi.nagoya-u.ac.jp/h/isets07/Contents/Session05/1018Vijay.pdf . . Weiss, R.F., 1974. Carbon dioxide in water and seawater: the solubility of a non-ideal gas. Mar. Chem. 2, 203 215. Wellinger, A., Murphy, J., Baxter, D., 2013. In: Hughes, S. (Ed.), The Biogas Handbook: Science, Production & Application. Woodhead Publishing Series in Energy. Woodhead Publishing, Philadelphia, PA, p. 2013. Wen, W.Y., Hung, J.H., 1970. Thermodynamics of hydrocarbon gases in aqueous tetraalkyl ammonium salt solutions. J. Phys. Chem. 74, 170 180. Wilhelm, E., Battlno, R., Wilcock, R.J., 1977. Low-pressure solubility of gases in liquid water. Chem. Rev. 77. Xiao, Y., Yuan, H., Pang, Y., Chen, S., Zhu, B., Zou, D., et al., 2014. CO2 removal from biogas by water washing system. Chin. J. Chem. Eng. 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 1

Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India, 2Department of Chemical Engineering, Indian Institute of Technology Delhi, New Delhi, India

Chapter outline 5.1 Introduction 5.2 Types of swing adsorption technologies 5.2.1 Temperature swing adsorption 5.2.2 Electric swing adsorption 5.2.3 Vacuum swing adsorption 5.2.4 Pressure swing adsorption 5.3 Parameters influencing pressure swing adsorption 5.3.1 Process performance indicators

5.1

93 95 95 97 97 98 99 100

5.3.2 Design parameters 5.3.3 Adsorbents 5.4 Adsorption isotherm 5.5 Adsorption kinetics 5.5.1 Molecular diffusion 5.5.2 Knudsen diffusivity 5.5.3 Poiseuille diffusion or viscous diffusion 5.6 Mathematical modeling 5.7 Conclusion and future perspectives References

101 103 107 109 110 110 110 112 113 113

Introduction

Discharge of carbon dioxide (CO2) into the environment accounts for 60% of the global warming caused by greenhouse gases. Over a 20-year period, methane is 84 times more potent than CO2 in terms 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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00005-2 © 2021 Elsevier Inc. All rights reserved.

93

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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 swingbased biogas purification technologies are progressively increasing for different applications. 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, chemical-free 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

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

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95

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.

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: G G G

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

TABLE 5.1 Efficiency of different swing adsorption technologies and adsorbents. S. No.

Technology

Adsorbent

Methane purity (%)

Cycle time (s)

References

1

VPSA

Carbon molecular sieves (CMS-3K)

.98%

400

Grande and Rodrigues (2007a,b)

2

PSA

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

.98%

800

Wu et al. (2015)

3

VPSA

Novel amine-based solid sorbent

.98%

900

Fujiki et al. (2017)

4

PSA

Carbon molecular sieve

#96%

840

Fatehi et al. (1995)

5

PSA

Zeolite 5A

$99%

500

Augelletti et al. (2017)

6

PSA

Zeolite 13X

$99%

800

Santos et al. (2011)

7

VPSA

Zeolite 13X (CECA)

90%

615

Cavenati et al. (2006)

8

PSA

Carbon molecular sieve (CMS)

$97.8%

990

Canevesi et al. (2019)

9

VPSA

Silica gel

98%

320

Shen et al. (2018)

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

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97

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

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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: G G G

Pressurization (with feed or raffinate product); High-pressure feed with raffinate withdrawal; Depressurization or blow down (cocurrent or countercurrent to the feed);

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99

Upgraded Gas

Purge Gas 4

Saturaon

Adsorpon

Feed

Biogas

3

2

HS Removal

Regeneraon

1

PSA Columns

CO -rich stream

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

G G G

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: 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.

100

5.3.1

PART | II Physiochemical upgrading systems

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 smallscale 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): Purity 5

NCH4 NCH4 1 NCO2

where, NCH4 and NCO2 are the number of moles of CH4 and CO2 respectively, that come out of the column in the elution step. Recovery 5

NCH4 FCH4

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. Productivity 5

FCH4 3 Recovery tcycle 3 Wtads

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: "
#
 γ21 Phigh γ γ Q Power 5 21  R  Tfeed  γ21 1000η Plow where γ 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 η 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.

Recent developments in pressure swing adsorption Chapter | 5

5.3.2

101

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

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PART | II Physiochemical upgrading systems

yet, implying that all the adsorbent in the bed remains to be saturated. Hence whatever feed is being passed 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 remember 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

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103

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 designing 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, carbonbased 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 ˚ ) so that methane diameter is less than the methane molecular diameter (3.8 A is not adsorbed inside the pores and at the same time it is greater than the CO2 ˚ ) so that the CO2 is adsorbed inside the pores. A molecule diameter (3.2 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 designing 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

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PART | II Physiochemical upgrading systems

´ lvarez-Gutie´rrez et al. (2018) during the initial stages of a PSA system. A 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 ´ lvarez-Gutie´rrez et al., 2018). (A 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. 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 liquidphase 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

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105

Activated carbons

Carbonbased nanomaterials

Carbonbased adsorbents

Carbon molecular sieves

Activated carbon fibers

FIGURE 5.3 Carbon-based adsorbents.

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

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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 Ca21 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 ˚ . Since molecules with diameters up to about 5 A ˚ can diameter of about 4.3 A ˚ sieve. The framework structures of X and penetrate, this is referred to as a 5 A 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 ˚ . Molecules with diameter up to about 8.5 A ˚ can a free diameter of up to 7.5 A 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 ˚ (Shokroo et al., 2016). free aperture of about 6 A 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) benzene-1,3,5-tricarboxylate] proved to be endowed with CO2 adsorption performances that are higher than those of typical adsorbents such as 13X zeolite.

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The distribution of pore size is classified into three major classes by size: G G G

˚; Micropores , 20 A ˚; Mesopores 20
500 A ˚ Macropores . 500 A.

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 ˚ is a reasonably common choice. general with relatively small molecules, 20 A 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 contact 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 contact 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 contact with the surface. q 5 kc 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.

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TABLE 5.2 Different types of adsorption isotherms. Isotherm Langmuir

Freundlich

Temkin

BrunauerEmmettTeller (BET)

Equation

qe 5

Plot

Q0 bCe 1 1 bCe

1 1 vs: qe Ce

1

qe 5 KF 3 Cen

logqe vs: logCe

qe 5

RT lnðAT Ce Þ b

lnqe vs: Ce

qe 5

qC C h s BET e ðCs 2 Ce Þ 1 1 ðCBET 2 1ÞðCCes Þ

Ce Ce vs: qs ðCs 2 Ce Þ Cs

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 A 1 S2SA. The equation derived from the previous assumptions takes the form: KLC q 5 qmax  1 1 KLC 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: q 5 KF Cn 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;

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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. The BET isotherm is of the form:
 1 c21 p 1 h  i5 1 p0 v p v c m 0 mc v p 21 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 5 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: G

G

G

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. 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

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Macropores

Bulk gas

Micropores

Solid surface

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

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: qffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi 1:86 3 1023 T 3=2 M11 1 M12 D5 pσ212 Ω 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: DA 5

dp pffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi 8RT=πMA 3

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

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contact 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: D5

Prp2 8μ

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: 1 Doverall

5

1 Dmolecular

1

1 Dmolecular

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: Deff 5

[ Doverall T

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: G G G G

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

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dynamic data. The LDF equation approximates the diffusion-controlled kinetics for many adsorption systems up to a satisfactory level. @q 15De  5 2 ðq 2 qÞ @t Rp

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). Overall material balance:

 @C 12ε @qCH4 @qCO2 @ðucÞ 1 50 1 ρs 1 @t ε @z @t @t

TABLE 5.3 Symbols used in the system equations of Section 5.6. Symbol

Unit

Meaning

C

mol/cm

qi

mol/g-adsorbent

3

ρs

gads/cm

Ci

3

Dax qi



3

Total concentration in gas phase Amount of gas ‘i’ adsorbed per unit gram of adsorbent Adsorbent density

mol/cm

Concentration in gas phase

2

cm /s

Axial dispersion coefficient

mol/g-adsorbent

Equilibrium adsorbed phase concentration

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Component material balance (two equations, one for CO2 and the other for CH4):
 @Ci 12ε @qi @2 C i @ðuci Þ 50 1 2 Dax 2 1 ρs ε @z @t @t @z Mass transfer rate (LDF equation) (one for each component): @qi  5 ki ðqi 2 qi Þ @t Equilibrium loading equation (Langmuir) (one for each component): 

qi 5

qsat;i Ki pi P Ki pi 11

Ergun equation (for pressure drop in the bed): @p 2 150μð12εÞ2 1:75ð1 2 εÞ 2 5 u1 u 2 3 @z dpε3 dp ε Ideal gas equation: P 5 CRT or yi 5 Ci RT

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, A., Idris, I., Shamsudin, I.K., Othman, M.R., 2019. Methane enrichment from high carbon dioxide content natural gas by pressure swing adsorption. J. Nat. Gas. Sci. Eng. 69, 102929.

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Ackley, M.W., Yang, R.T., 1990. Kinetic separation by pressure swing adsorption: method of characteristics model. AIChE J. 36 (8), 1229
1238. ´ lvarez-Gutie´rrez, N., Gil, M.V., Rubiera, F., Pevida, C., 2018. Simplistic approach for prelimiA nary screening of potential carbon adsorbents for CO2 separation from biogas. J. CO2 Utilization 28, 207
215. Arya, A., Divekar, S., Rawat, R., Gupta, P., Garg, M.O., Dasgupta, S., et al., 2015. Upgrading biogas at low pressure by vacuum swing adsorption. Ind. & Eng. Chem. Res. 54 (1), 404
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196. Fatehi, A.I., Loughlin, K.F., Hassan, M.M., 1995. Separation of methane—nitrogen mixtures by pressure swing adsorption using a carbon molecular sieve. Gas. Sep. Purif. 9 (3), 199
204. Fujiki, J., Chowdhury, F.A., Yamada, H., Yogo, K., 2017. Highly efficient post-combustion CO2 capture by low-temperature steam-aided vacuum swing adsorption using a novel polyaminebased solid sorbent. Chem. Eng. J. 307, 273
282. Grande, C.A., Rodrigues, A.E., 2007a. Biogas to fuel by vacuum pressure swing adsorption I. Behavior of equilibrium and kinetic-based adsorbents. Ind. & Eng. Chem. Res. 46 (13), 4595
4605. Grande, C.A., Rodrigues, A.E., 2007b. Layered vacuum pressure-swing adsorption for biogas upgrading. Ind. Eng. Chem. Res. 46 (23), 7844
7848. Haghpanah, R., Majumder, A., Nilam, R., Rajendran, A., Farooq, S., Karimi, I.A., et al., 2013. Multiobjective optimization of a four-step adsorption process for postcombustion CO2 capture via finite volume simulation. Ind. & Eng. Chem. Res. 52 (11), 4249
4265. Heck, H.H., Hall, M.L., dos Santos, R., Tomadakis, M.M., 2018. Pressure swing adsorption separation of H2S/CO2/CH4 gas mixtures with molecular sieves 4A, 5A, and 13X. Sep. Sci. Technol. 53 (10), 1490
1497. Ho, Y.S., 2006. Review of second-order models for adsorption systems. J. Hazard. Mater. 136 (3), 681
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43. Li, Z., Xiao, G., Yang, Q., Xiao, Y., Zhong, C., 2014. Computational exploration of metal
organic frameworks for CO2/CH4 separation via temperature swing adsorption. Chem. Eng. Sci. 120, 59
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Mosca, A., 2009. Structured zeolite adsorbents for PSA applications (Doctoral dissertation, Lulea˚tekniskauniversitet). Nikolaidis, G.N., Kikkinides, E.S., Georgiadis, M.C., 2017. 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. 343. Ntiamoah, A., Ling, J., Xiao, P., Webley, P.A., Zhai, Y., 2016. CO2 capture by temperature swing adsorption: use of hot CO2-rich gas for regeneration. Ind. & Eng. Chem. Res. 55 (3), 703
713. Petkovska, M., Antov-Bozalo, D., Markovic, A., Sullivan, P., 2007. Multiphysics modeling of electric-swing adsorption system with in-vessel condensation. Adsorption 13 (3-4), 357
372. Pirngruber, G.D., Hamon, L., Bourrelly, S., Llewellyn, P.L., Lenoir, E., Guillerm, V., et al., 2012. A method for screening the potential of MOFs as CO2 adsorbents in pressure swing adsorption processes. ChemSusChem 5 (4), 762
776. Qiu, H., Lv, L., Pan, B.C., Zhang, Q.J., Zhang, W.M., Zhang, Q.X., 2009. Critical review in adsorption kinetic models. J. Zhejiang University-Science A 10 (5), 716
724. Reid, C.R., Thomas, K.M., 1999. Adsorption of gases on a carbon molecular sieve used for air separation: linear adsorptives as probes for kinetic selectivity. Langmuir 15 (9), 3206
3218. Ribeiro, R.P.P.L., Grande, C.A., Rodrigues, A.E., 2014. Electric swing adsorption for gas separation and purification: a review. Sep. Sci. Technol. 49 (13), 1985
2002. Riboldi, L., Bolland, O., 2017. Overview on pressure swing adsorption (PSA) as CO2 capture technology: state-of-the-art, limits and potentials. Energy Procedia 114, 2390
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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

Chapter outline 6.1 Introduction: how the basic membrane processes for gas separation have evolved 117 6.2 Basic terms 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 absorptionmembrane 145 6.6.3 Microbial conversion of CO2 to CH4 on a membrane diffuser 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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00006-4 © 2021 Elsevier Inc. All rights reserved.

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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. 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

FIGURE 6.1 Timeline for the development of gas separation membranes.

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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 Roadmap 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 water-scrubbing, 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): G G G G

G

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 by-products; 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.

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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 terms 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. In this context, the basic equations that describe the material flow in the process are: Flow balance: FF 5 FR 1 FP

ð6:1Þ

Mass balance for component i: FF  Ci 5 FR  CiR 1 FP  CiP

ð6:2Þ

And for design of the membrane processes, other relevant parameters are: The ratio of feed pressure to permeate pressure: ψ5

PF PP

ð6:3Þ

And the ratio of permeate to feed flow (or stage-cut): θ5

FP FF

ð6:4Þ

Effective separation takes place when some gas components pass through the membrane barrier quicker than others. This process can be described by

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.

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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. The performance of the membrane for separation of a given binary gas mixture of components A and B is typically expressed as follows:     PA DA SA αA=B 5 5  ð6:5Þ PB DB SB 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: αA=B 5

yA =yB xA =xB

ð6:6Þ

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 pressure-normalized 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: 1 GPU 5 1026 cm3 ðSTPÞ=cm2 s cm Hg 1 GPU 5 7:6 3 1029 m3 ðSTPÞ=m2 s kPa

ð6:7Þ

In the SI system, permeance is expressed in the following unit: ½ P 5

mol m s Pa

ð6:8Þ

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. 96103.

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However, another widely used and accepted unit for permeability is the Barrer: 1 Barrer 5 10210 cm3 ðSTPÞ cm=½cm2 s ðcm HgÞ

ð6:9Þ

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 terms 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 (Ha¨gg 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 passage, 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.

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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. Most commercial membranes are made as asymmetric membranes consisting of a thick porous layer or support, 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 support 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.

TABLE 6.1 Common polymer membrane materials and their gas transport properties for the CO2/CH4 binary gas mixture. Polymer

P (CO2) Barrer

α (CO2/CH4)

Reference

Cellulose acetate (CA)

8.9

2025

Spillman (1995)

Polyimide (PI)

65110

2515

Ismail et al. (2015)

25368

13.134.5

A´lvarez et al. (2020)

Polyamide (PA)

—

12.814.4

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)

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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 (503000 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 support 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 support 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.

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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 passes 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 low-permeate 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 join 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

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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 gasliquid membrane contactors, 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. Employing membranes rather than packed columns for gasliquid contacting and absorption offers several advantages like independence from flooding, weeping, and over-loading, hence giving a better freedom of choice of mass-transfer coefficient and contact 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 gasliquid membrane contactors for CO2 absorption would require further advancements in the following topics: 1. Improvement of present microporous membranes in terms of antiwettability, and thermal and chemical resistance;

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

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2. Development of novel absorbents with high separation effectivity, good compatibility with the membrane, and with the possibility of costefficient regeneration; 3. Development of novel membrane configurations and modules that enhance mass transfer. Gasliquid 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 L-argininate 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 support. 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 contactors that combine the advantages of both unit operations.

6.3 6.3.1

Membrane materials and structures 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 accounts for preferential passage 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 passage 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

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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 passage of a over b, whereas the higher condensability of b would preferentially allow passage of b over a (Drioli and Giorno, 2009). 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): G

G

G

G

Air separation (production of technical-grade nitrogen-or oxygenenriched 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. 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.

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.

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TABLE 6.2 Principal membrane companies and membrane materials used in gas separation (Baker, 2002; Bernardo et al., 2009). 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 a

Kryogenmash

Poly(vinyltrimethylsilane), tetrabromopolycarbonate

Air Liquid

Ethyl cellulose

OPW Vaposaver

Poly(trimethylsilyl propyne)

a

Producer of membrane modules and installations only.

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 Si-substituted 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.

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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 terms 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, CH:3 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). 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

TABLE 6.3 Properties of 6FDA polyimide monomers.

R1

R2

P(O2), Barrer

α (O2/N2)

Substituent volume, A˚3

H

H

2.8

5.65

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

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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. 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 mass-transfer 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 (Ha¨gg 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.

TABLE 6.4 Effects of the connector groups in 6FDA polyimides of general formula.

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), 10591064.

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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 support 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 ˚ (Baker, 2004). Both can withstand can contain pores in the range of 38 A high temperatures and have good chemical resistance. However, they are expensive and brittle. Zeolites have been used dispersed in polymers in the form of mixed-matrix membranes (Pechar et al., 2006; Bastani et al., 2013) or as support for gasliquid membrane contactors (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 Ha¨gg and He (2011), where it is stated that producing a high-performance 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. 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 (Ha¨gg 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

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FIGURE 6.9 Important steps in the manufacture of CMS membranes. Adapted from Ha¨gg, M.-B., He X., 2011. Chapter 15. Carbon Molecular Sieve Membranes for Gas Separation. 162191.

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.

of different additives dependent on the precursor polymer chosen (Ha¨gg and He, 2011; Prangsgaard Andersen, 2016) (Fig. 6.10). The Israeli company Carbon Membranes Ltd. made full-scale CMS membrane modules in 19952001 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

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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.% CMSloaded 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 support in MMMs or as a selective layer on composite membranes (Esposito et al., 2015).

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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 pass 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 (Vrbov´a and Ciahotny´, 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 6.4.1

Theory of transport in gas separation on membranes Transport through rubbery polymers

Rubbery polymers act in such a way that gases get absorbed in the membrane material following Henry’s law: ½Ci liq 5 Hi  pi

ð6:10Þ

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 pass from one side of the membrane to the other side driven by a concentration gradient. In the case

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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: ϕ 5 PðΔp=lÞ

ð6:11Þ

where ϕ is the flux or steady-state permeation rate, P is the permeability coefficient, Δp 5 P2 2 P1 is the pressure difference between the two sides of the membrane, and l is the thickness of the membrane. In a fundamental formulation, the basis for permeation is phrased by Fick’s first law at steady state at any point inside the membrane:     Dloc dC ϕ5 2 ð6:12Þ  dx 12w 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: ϕi 5 Di Δ½Ci =l 5 Di  Hi  Δpi =l 5 Di  Si  Δpi =l

ð6:13Þ

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: P5D  S

ð6:14Þ

P, D, and S values in polymers can be determined using different strategies and these methods have been thoroughly reviewed by Felder and

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

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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: ½Ci liq 5 Ki pi 1

½Ci L bpi 1 1 bpi

ð6:15Þ

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. Simple homogeneous swelling of rubbery polymers by compatible penetrants tends to produce convex isotherms contrary to Fig. 6.12 (2). More

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

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).

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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. 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

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)].

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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 6FDA-TMODA/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

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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): G

G

G

G

Flow is low to moderate: 50005,000,000 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 [1050 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. 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 ˇ ak et al. (2018) reported a two-stage biogas this process. On the other hand, Z´ 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

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

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

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.

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

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production of the order of 35,00075,000 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 microbiological 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 cryogenicmembrane 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

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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, 194202.

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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 cryogenic-membrane 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 flash-membrane systems.

6.6.2 Biogas upgrading by absorption and hybrid absorptionmembrane 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. In this arrangement, the solvent is supplied together with the feed gas to the high-pressure side of the supported 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

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FIGURE 6.17 Representation of bulk flow liquid membrane.

FIGURE 6.18 Hybrid membrane-absorption flowsheet.

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 absorber-stripper 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. 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 9699 vol.%. Also, in this case, studies of hybrid membrane-absorption processes are limited but the results discussed highlighted how membrane-aided processes

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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 passed 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

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).

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

requirement of low water and H2S content. The process is described in more detail in 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 terms of upgrading costs using Aspen Plus. Investment costs are accounted 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 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 waterscrubbing process (Weidner et al., 2008).

6.6.3

Microbial conversion of CO2 to CH4 on a membrane diffuser

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

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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 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: CO2 1 4H2 -CH4 1 2H2 O without recycling of the mixed gases.

ð6:16Þ

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

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

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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 1502000 N m3/h, data prices in the range of 0.140.075 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 0.140.075 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.”

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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.

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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

Chapter outline 7.1 Introduction 159 7.2 Cryogenic biogas upgrading 160 7.2.1 Cryogenic distillation 162 7.2.2 Cryogenic packedbed 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

7.1

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

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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00007-6 © 2021 Elsevier Inc. All rights reserved.

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digestion of waste (Angelidaki et al., 2018). To obtain a 100% green source of energy, CO2 needs to be removed before biogas utilization (BaenaMoreno 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 (Baena-Moreno 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

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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. 2. 3. 4. 5.

Indirect contact between gas and chemicals; Production of high-purity CO2 as a secondary product; Good separation of all the components present in biogas; Production of liquefied biogas (LBG); 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. Technologies based on cryogenic separation processes are more energyefficient 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).

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/biogasopwaardering. (accessed 16.11.19).

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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.

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 (2164 C for CH4 and 278 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). 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 240 C, keeping the pressure value;

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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. 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 terms of specific energy demand. These new findings could make cryogenic distillation applicable and competitive in comparison with traditional biogas upgrading approaches.

7.2.2

Cryogenic packed-bed technology

Cryogenic packed bed (CPB) 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 CPB the bed is cooled to temperatures lower than 2100 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.

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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.

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

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TABLE 7.1 Comparison between the current process and the traditional biogas upgrading approaches in terms of specific energy demand. Biogas upgrading techniques

Specific energy demand (kWh/Nm3 cleaned gas)

Reference

Cryogenic distillation (2 distillation columns)

0.5 0.6

Yousef et al. (2018)

Water scrubbing

0.45 0.92

Persson Owe Jo¨nsson Arthur Wellinger et al. (2006)

Physical absorption

0.5 0.7

Bauer et al. (2013)

Chemical absorption

0.4 0.45

Nilsson (2001)

Pressure swing adsorption

0.3 1

Persson Owe Jo¨nsson Arthur Wellinger et al. (2006)

Membrane technology

0.25 0.42

Bauer et al. (2013)

Typical cryogenic separation (one distillation column)

0.8 1.55

Ali et al. (2010)

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.

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.

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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, CPB 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. 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 terms 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

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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.

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

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FIGURE 7.6 Cryogenic-absorption combination scheme for biogas upgrading. Adapted from ´ M., Vega F., Vilches L.F., Navarrete B., Zhang Z., 2019f. Baena-Moreno F.M., Rodr´ıguez-Galan Biogas upgrading by cryogenic techniques. Environ. Chem. Lett. 17, https://doi.org/10.1007/ s10311-019-00872-2.

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 220 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.

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

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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.

synergized with a low-temperature 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 terms 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).

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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 cryogenic-membrane processes can be seen in Section 7.4.

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

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.

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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.

subjected to 255 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.

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 Ha¨gg, 2010; Song et al., 2019). Fig. 7.10 represents the two-stage 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. 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 three-stage configuration, achieving CO2/N2 selectivity over 50. Further studies would be needed to estimate these results for biogas upgrading.

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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: stateof-the-art developments and current challenges. Renew. Sustain. Energy Rev.101:265 278. https://doi.org/10.1016/j.rser.2018.11.018.

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

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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 membranecryogenic 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 220 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/BPDADAM 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

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

TABLE 7.2 Summary of studies from membrane-cryogenic hybrid processes. Flue gas source

CO2 concentration (%)

Product purity (%)

Energy consumption (MJ/t CO2)

Reference

Gas combustion (coal)

13.9

95

0.94 1.08

Mat and Lipscomb (2019)

Gas combustion (coal)

15.0

99.8

1.70

Song et al. (2017)

Gas combustion (coal)

13.5

98.3

1.22

Zhao et al. (2014)

Biogas

99.9 (CH4)

Scholz et al. (2013)

Gas combustion (coal)

14.0

94.1

1.25

Shafiee et al. (2017)

Gas combustion (coal)

13.7

95

1.11

Zhang et al. (2014)

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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 passes 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 terms 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 terms 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 2110 C to 2 88.3 C, which implied a relevant reduction of the cooling energy requirements for the whole process. The optimized operation of this novel integration membranecryogenic 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

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 membrane-cryogenic hybrid CO2 capture by process optimization. Energy. 124, https://doi.org/10.1016/j.energy.2017.02.054.

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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 250 C and 235 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 terms 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

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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 terms 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 (2120 C) to 2 bar (2160 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 h/kWh and for OPEX around 4.80 7.10 h/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 h/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.

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TABLE 7.3 Comparisons of the described technologies. Technology

Advantages

Cryogenic distillation

G

G

G

Cryogenic packed-bed technology

G

G

G

G

Cryogenic technology with absorption

G

G

G

Cryogenic technology with adsorption

G

G

Disadvantages

High quality of the product streams (97.2% CH4; 99.3% CO2) New studies show a reduction in the specific energy demand (0.5 0.6 kWh/Nm3 of cleaned gas) Competitive technology with other traditional biogas upgrading techniques (water scrubbing, physical absorption, chemical absorption, and pressure swing adsorption)

G

Technology applied to a wide range of other gas separations CO2 capture is not required for biogas treatment. Maximum CO2 removal is guaranteed There is no problem of frosting along the system Simple construction of the beds

G

Utilization of a solvent at temperatures below the usual working ones The use of ammonia as an absorbent is cheap, has high availability and low energy regeneration Biomethane is obtained in the absorption stage together with the CO2 liquid stream

G

Typical absorbents employed (NaOH, KOH, or MEA) do not show a high performance

High potential to be combined with other technologies Highly researched in recent years

G

Purity of product around 88.9% of CO2 recovery High consumption of energy Low economy performance of the process

G

G

G

G

G

G

Biogas must achieve very low temperature to be separated (2164 C for CH4 and 278 C for CO2 at 1 bar) The minority of impurities in the biogas must be removed before the separation stage Presence of CO2 freezing out along the system, causing frosting, and increasing energy consumption

Biogas is pressured higher than 40 bars The packaged-bed must be refrigerated before operating Purity of product is lower than other techniques (around 94%)

(Continued )

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TABLE 7.3 (Continued) Technology

Advantages

Cryogenic technology with membranes

G

G

G

G

Cryogenic technology with hydrate technologies

G

G

G

7.7

Disadvantages

The most promising strategy in the field of cryogenic biogas upgrading Development of simple membranes and environmentally friendly materials The energy consumption of the biogas upgrading system can be reduced by analyzing different configurations of the system High purity in the final stream

G

Promising alternative extensively studied by experts High-purity CO2 streams can be obtained. Allows the separation of contaminants in the biogas in the first stage

G

G

G

G

G

G

Low resilience under longterm operations Low CO2 separation targets using a single-membrane separation technique Degradation in the presence of pollutant, mainly oxides from nitrogen and sulfur Low CO2 selectivity and low CO2 permeability with materials of current membranes It is necessary to search for new materials, cooling sources, and optimized configuration processes High pressure in the process High-pressure water consumption

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 CPB 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 hydrate-cryogenic methods has been carried out. Among the presented integrations, cryogenic-membrane technology is highlighted as the energy consumption can be considerably reduced without

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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: kWh 3:6MJ 1 Nm3 YCH4 1 kmol YCH4 3 3 3 3 5 3:65 3 3 1kWh 0:044 tCO2 22:4 kmol YCO4 YCO4 Nm

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.

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

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

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Chapter 8

Power-to-gas for methanation Anirudh Bhanu Teja Nelabhotla1, Deepak Pant2 and Carlos Dinamarca1 1

Department of Process Energy and Environmental Technology, University of South-Eastern Norway, Porsgrunn, Norway, 2Separation and Conversion Technology, Flemish Institute for Technological Research (VITO), Mol, Belgium

Chapter outline 8.1 Introduction 187 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

8.1

8.2.6 Three-phase reactor 8.2.7 Micro(channel) reactors 8.3 Bioelectrochemical methanation 8.3.1 Direct electron transfer 8.3.2 Biocathodes 8.3.3 Reactor configurations 8.4 Challenges and future prospects References

202 203 205 207 209 210 211 213

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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00008-8 © 2021 Elsevier Inc. All rights reserved.

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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). PtG technology provides an opportunity to reintroduce carbon dioxide into the energy cycle by reducing it to various chemicals or gases. This in turn supports 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. TABLE 8.1 Storing electricity generated intermittently by solar and wind sources (Specht et al., 2010). 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

Power-to-gas for methanation Chapter | 8

189

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 Germany, 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 h/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 terms 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 terms of research and large-scale production prototypes. Chemical methanation involves high-temperature and highpressure 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.

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

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TABLE 8.2 Primary differences between chemical and biological methanation. Parameter

Chemical


Bioelectrochemical


Temperature

200 C800 C

20
C50
C

Pressure

20100 bars

13 bar

Catalysts

Expensive nobel metals

Microorganisms

Reactor footprint

Small

Small

Operational flexibility

Mediumrigid

Flexible

Capital expenditure

High

Medium

Operational expenditure

High

Medium

Production efficiencies

High

Medium

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, threephase, and other reactor designs being developed. Electrolysis reactions (Rabaey and Rozendal, 2010) include the following: Anode reaction 2H2 O-O2 1 4H1 1 4e2

E
510:81 V vs: NHE

ð8:1Þ

E
5 20:414 V vs: NHE

ð8:2Þ

Cathode reaction 2H1 1 2e2 -H2

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

8.2.1

4H2 1CO2 -CH4 1 2H2 O

ΔG 5 2165 kJ=mol

ð8:3Þ

3H2 1 CO-CH4 1H2 O

ΔG 5 2206 kJ=mol

ð8:4Þ

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 (Ursu´a et al., 2012). However, the process is very energy demanding with high capital investment, operation, and maintenance costs (Manabe et al., 2013).

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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 2040 wt.%. AWE is operated at approximately 7090
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 2H2 O 1 2e2 -H2 1 2OH2

ð8:5Þ

M1H2 O 1 2e2 -MHadsorbed 1OH2

ð8:6Þ

2MHadsorbed -2M1H2

ð8:7Þ

H2 O 1 MHadsorbed 1 e2 -M1H2 1OH2

ð8:8Þ

Tafel reaction

Or Heyrovsky reaction

Anode reaction 4OH2 -O2 12H2 O 1 4e2

ð8:9Þ

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

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

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 5 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; Stoji´c 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. 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 terms of ohmic losses due to the distance between electrodes.

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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

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

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PART | II Physiochemical upgrading systems

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
C100
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, 25 times more than AWE (Michishita et al., 2008; Takenaka, 1991). 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.

8.2.2.2 Reactor configurations The main element of a PEM reactor is the membrane electrode assembly (MEA), where the electrodes, gas diffuser, 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 (20300 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).

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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.

8.2.2.3 Recent developments 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, Co-doped FeS2/CNTs, WO2/C nanowires, and CoFe nanoalloys encapsulated in N-doped graphene (Shiva Kumar and Himabindu, 2019). Recent studies showed improved HER performance with increased current density using synthesized RuS2@MoS2 nanoparticlenanosheet 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). N-doped 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

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PART | II Physiochemical upgrading systems

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

8.2.3.1 Design and concept Solid oxide electrolyzers (SOEs) are the most recent development in electrolyzer technology that work on the principle of anion exchange in the form of O22 (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 largescale operation. SOE is operated under high-pressure and high-temperature (500
C850
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

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

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197

cathode go through the solid electrolyte to the anode, where they recombine, according to Eq. (8.10), to form oxygen (Ursu´a et al., 2012). Electrolysis reactions include the following: Anode reaction 2 2O2 2 -O2 1 2e

E
510:81 V vs: NHE

ð8:10Þ

Cathode reaction H2 O 1 2e2 -H2 1 O2 2

E
5 20:414 V vs: NHE

ð8:11Þ

8.2.3.2 Reactor configurations The SOE demands porous electrodes to maximize the contact 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 calcia-stabilized 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). 8.2.3.3 Recent developments Most recent developments in the SOE in terms of power-to-gas have focused on reactor configuration and strategizing the integration with methanation reactors. Some studies have carried out technoeconomic analyses in terms 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

198

PART | II Physiochemical upgrading systems

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-to-methane second to power-to-hydrogen. However, the study concluded that solid oxide electrolysis will be most suitable for PtG in terms 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

8.2.4.1 Design and concept 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

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

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199

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).

8.2.4.2 Reactor configurations 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
C800
C and 2580 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 fixed-bed 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). 8.2.4.3 Recent developments 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

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PART | II Physiochemical upgrading systems

can be reactivated after being exposed to long hours of nonreacting carbonaceous reactants, that is, CO2 or CO. The reactor design and materials must support frequent starting and stopping of methanation without start-up delays and inefficiencies (Ro¨nsch 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

8.2.5.1 Design and concept 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 5 GF), (2) circulating (CS . GF), and (3) transport (CS , GF). The efficient heat transfer and mixing allows the operating temperatures of fluidized bed reactors to be lower than fixed-bed reactors, that is, around 200
C500
C, and operating pressures in the range of 2060 bars (Cocco et al., 2014). 8.2.5.2 Reactor configurations 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 (Go¨tz 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

Power-to-gas for methanation Chapter | 8

201

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 (Containerbased 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). 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 support 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).

8.2.5.3 Recent developments The bubbling-type reactor has taken the center stage in terms of the development of the fluidized bed methanation reactor research. In bubbling-type fluidization, the reactant gas passes 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

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

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PART | II Physiochemical upgrading systems

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 twofluid 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

8.2.6.1 Design and concept 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 gasliquid mass transfer limitations and also decomposition and evaporation of the heat transfer liquid (Go¨tz et al., 2016). 8.2.6.2 Reactor configurations 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

Power-to-gas for methanation Chapter | 8

203

FIGURE 8.6 Schematic diagram of the three-phase reactor (Go¨tz et al., 2014).

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- propyl-piperidinium bis(trifluoromethylsulfonyl) imide [PMPip] [BTA], and 1-ethyl-3-methylimidazolium trifluoromethanesulfonate [EMIM] [Tf]} that showed positive results on a 3PM operation (Lefebvre et al., 2018; Go¨tz et al., 2013).

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 areato-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 supports 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

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PART | II Physiochemical upgrading systems

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
C500
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 support 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). 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

FIGURE 8.7 Engineering (Brooks et al., 2007).

drawing

of

the

Sabatier

microchannel

reactor

assembly

Power-to-gas for methanation Chapter | 8

205

complete removal and reduction of CO to CH4 with Ru/SiO2 combination catalyst coated directly on to the stainless steel microchannel reactors (Go¨rke 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 contact time of milliseconds (Liu et al., 2012). The authors created novel metalceramics complex substrate that was employed as the catalyst support, 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 Eerten-Jansen 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 20.414 V, whereas the microorganisms, as catalysts, generate a cathode potential of about 20.300 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

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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).

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). 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 support determines the electrochemical cell efficiency. Indirect electron transfer: CH3 COOH 1 2H2 O-2CO2 1 8H1 1 8 e2 8H1 1 8e2 -4H2

E
5 20:280 V vs: NHE ð8:12Þ

E
5 20:414 V vs: NHE

4H2 1 CO2 -CH4 1 2H2 O

ΔG 5 2165 kJ=mol

ð8:13Þ ð8:14Þ

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Direct electron transfer: 4H2 O-2O2 1 8H1 1 8e2 CO2 1 8H1 1 8e2 -CH4 1 2H2 O

E
510:810 V vs: NHE

ð8:15Þ

E
5 2 0:244 V vs: NHE

ð8:16Þ

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.

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 contact, electrode

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).

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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 verify 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) 5 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).

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8.3.2

209

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 20.65 to 20.75 V (vs. SHE), methane, hydrogen, and acetic acid at 20.75 V (vs. SHE), and methane and acetic acid at 20.95 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).

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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 in-depth 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;

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

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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.

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.

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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.

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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

Chapter outline 9.1 Introduction 9.2 Faradaic and energy efficiency 9.3 Electroreduction of CO2 9.3.1 Basic considerations 9.3.2 Reactor and process design 9.4 Electrochemical oxidation of H2S 9.4.1 Basic considerations 9.4.2 Reactor and process design

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9.5 Biogas upgrading approach and its challenges 9.5.1 CO2 electroreduction 9.5.2 H2S oxidation 9.5.3 Biogas and scale-up approaches 9.6 Concluding remarks and perspectives Acknowledgments References

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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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00009-X © 2021 Elsevier Inc. All rights reserved.

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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 (Cl2), 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 Germany, 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 terms of metabolic pathways are used: chemoautotrophy

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225

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).

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PART | II Physiochemical upgrading systems

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. FE ð%Þ 5 EE ð%Þ 5

znF U100 Q

ð9:1Þ

Et U100 E

ð9:2Þ

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 5 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 hightemperature 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/H2O-liquid 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; S´anchez 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

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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.

most exploited aspects of the CO2R process and its focus on biogas upgrading will be presented.

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: xCO2 1 nH 1 1 ne2 -product 1 yH2 O

ð9:3Þ

xCO 1 nH 1 1 ne2 -product 1 yH2 O

ð9:4Þ

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PART | II Physiochemical upgrading systems

Following Eqs. (9.1) and (9.2), the CO2R process is considered as a multiple-step 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: 2H2 O-4H 1 1 4e2 1 O2 2H 1 1 2e2 -H2

ðOER
anodeÞ

ðHER
cathodeÞ

ð9:5Þ ð9:6Þ

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. 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: 1 ð9:7Þ CO2 -CO 1 O2 ðcathodeÞ 2 Steam electrolysis: H2 O-H2 1

1 O2 2

ðcathodeÞ

ð9:8Þ

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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.

Reverse water-gas shift reaction (RWGS): H2 1 CO2 -CO 1 H2 O

ðcathodeÞ

ð9:9Þ

Oxygen formation: O2 2 -2e2 1

1 O2 2

ðanodeÞ

ð9:10Þ

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): 8e2 1 CO2 1 8H 1 -CH4 1 H2 O

ðcathodeÞ

ð9:11Þ

The drawback of solid-oxide devices in terms 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

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PART | II Physiochemical upgrading systems

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 co-electrolysis 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.

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.

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).

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231

The electrodes are porous, most commonly ceramic composite materials made of nickel and yttria-stabilized zirconia (YSZ). The solid electrolyte can be either oxygen ion-conducting or proton-conducting. In the first case, water and CO2 are reduced in the cathode and oxygen ions pass 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 lab-scale 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 CO322, which reduces the availability of the dissolved CO2 in the electrolyte. Nevertheless, alkaline conditions favor CO2R over competing HER which occurs at low pH (Ko¨nig 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 sandwich-type. 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, continuousflow 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,

232

PART | II Physiochemical upgrading systems

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 designing 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) support 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 (S´anchez 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). According to this summary, the highest Faradaic efficiencies can be observed when CO is the target product of the CO2R process. An efficiency

TABLE 9.1 Operating conditions and product performances recorded for various designs used for electroreduction of CO2R. Reactor type

H-type cell

Conventional H-type cell

Sandwichtype cell

Electrocatalyst

Substrate

Electrolyte type

Electrolyte composition

Membrane type

Temperature/ pressure

Current density (mA/cm2)

Main product (FE)

A-Ni-NG (Yang et al., 2018)

Glassy carbon

CO2 sat. solution

0.5 M KHCO3

Nafion 117 (CEM)

RT/ ambient

22 (20.61 V vs. RHE)

CO (97%)

NS-C (Pan et al., 2019)

Carbon paper

CO2 sat. solution

0.1 M KHCO3

Nafion 115 (CEM)

RT/ ambient

2.86 (20.49 V vs. RHE)

CO (92%)

Np-Ag (Lu et al., 2014)

GDL

CO2 sat. solution

0.1 M KOH

AEM

RT/ ambient

10 (20.5 V vs. RHE)

CO (92%)

OD-Cu (Kas et al., 2014)

Cu foil

CO2 sat. solution

0.1 M KHCO3

Nafion 112 (CEM)

RT/ ambient

36 (21.1 V vs. RHE)

C2H4 (33%)

Cu nanocube (Jiang et al., 2018)

Cu foil

CO2 sat. solution

0.25 M KHCO3

Nafion 117 (CEM)

RT/ ambient

68 (20.96 V vs. RHE)

C2 (over 60%)

PdPt/C (Kortlever et al., 2015a)

Graphite electrode

CO2 sat. solution

0.1 M K2HPO4/ 0.1 M KH2PO4

Nafion 115 (CEM)

RT/ ambient

5 (20.4 V vs. RHE)

HCOOH (88%)

Cu foil (Kuhl et al., 2012)

Cu foil

CO2 sat. solution

0.1 M KHCO3

AEM

RT/ ambient

6 (21.05 V vs. RHE)

C1 and C2 (over 75%)

Cu NPs (Manthiram et al., 2014)

Glassy carbon plate

CO2 sat. solution

0.1 M NaHCO3

AEM

RT/ ambient

12 (21.35 V vs. RHE)

CH4 (80%)

AN-Cu (Lee et al., 2018)

Cu foil

CO2 sat. solution

0.1 M KHCO3

AEM

RT/ ambient

20 (21.08 V vs. RHE)

C2H4 (38.1%)

(Continued )

TABLE 9.1 (Continued) Reactor type

Pressurized reactor Flow cell

PEM

Electrocatalyst

Substrate

Electrolyte type

Electrolyte composition

Membrane type

Temperature/ pressure

Current density (mA/cm2)

Main product (FE)

Ag foil (Hatsukade et al., 2014)

Ag foil

CO2 sat. solution

0.1 M KHCO3

AEM

RT/ ambient

4 (21.02 V vs. RHE)

CO (over 90%)

Cu NPs (Kas et al., 2015)

Cu plate

CO2 sat. solution

0.1B0.5 M KHCO3

Nafion 112 (CEM)

RT/1
9 atm

(21.8 V vs. Ag/AgCl).

C2H4 1 CH4 (45%)

In/Pb (Narayanan et al., 2011)

Carbon paper

CO2 sat. solution

1 M NaHCO3

Nafion 115/ AEM

RT/ ambient

40

HCOOH (80%)

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

GDL

H2 gas 1

/

CEM/AEM

RT/ ambient

/

CO

1 M NaOH

BPM

RT/ ambient

200

CO (50%)

1 M KOH

AEM

RT/ ambient

150

CO (over 95%)

10 mM KHCO3

AEM

RT/ ambient

B200

CO (over 90%)

0.1 M KHCO3

Nafion 117 (CEM)

RT/ ambient

/

C 2 H4 (91.1%)

0.5 M KHCO3

Nafion 117 (CEM)

RT/ ambient

/

CH3OH (25.6%)

CO2 gas Ag (Salvatore et al., 2018)

GDL

Solution 1 CO2 gas

CoPc (Ren et al., 2019)

GDL

Solution 1

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

Carbon paper

Solution 1

Cu oxides/ZnO (MerinoGarcia et al., 2019)

Carbon paper

Solution 1

CO2 gas

CO2 gas

CO2 gas Cu2O/ZnO (Albo et al., 2017)

Carbon paper

Solution 1 CO2 gas

Microfluidic

SOEC

RuPd/Sn (Whipple et al., 2010)

GDE

Solution 1 CO2 gas

0.5 M KCl 1 1 M HCl or 1 M KOH

/

RT/ ambient

B100 (20.55 V vs. RHE)

HCOOH (89%)

Ag (Rosen et al., 2011)

GDL

Solution 1 CO2 gas

18 mol% EMIM BF4 in water

Nafion 117 (CEM)

RT/ ambient

/

CO (96%)

CuNP (Ma et al., 2016)

GDL

Solution 1 CO2 gas

1 M KOH

AEM

RT/ ambient

B200 ( 2 0.7 V vs. RHE)

C2H4 (46%)

Graphite/carbon NPs/Cu/ PTFE (Dinh et al., 2018)

GDL

Solution 1 CO2 gas

7 M KOH

AEM

RT/ ambient

B100 ( 2 0.55 V vs. RHE)

C2H4 (70%)

La0.80Sr0.20Sc0.05Mn0.95O3- Gd0.20Ce0.80O1.95 (Yu et al., 2018)

The electrolyte tube

Air 1 CO2 gas

(YSZ) electrolyte tubes

/

800 C

230

CO

Ni/YSZ (Ebbesen et al., 2012)

The electrolyte tube

CO2
H2 or CO2
H2O
CO mixtures

(YSZ) electrolyte tubes

/

750 C/850 C

/

CO

Cu sheet (Clark et al., 2015)

Cu sheet

CO2 sat. solution

0.05 M K2CO3

Nafion 110 (CEM)

RT/ambient

/

C1 and C2

Cu/Ag (Clark and Bell, 2018)

PTFE membrane

CO2 sat. solution

0.05 M Cs2CO3

AEM

RT/ambient

/

C1 and C2/ CO

d

DEMS cell

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.

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PART | II Physiochemical upgrading systems

as high as 97%, along with a low overpotential of 20.61 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 sandwich-type 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/ZnO-based electrodes, which revealed excellent FE equal to 91.1% and a production rate of 487.9 6 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 6 136 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 terms 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.

Electrochemical approach for biogas upgrading Chapter | 9

9.4.1

237

Basic considerations

Hydrogen sulfide is a weak acid and in aqueous solutions dissociates to H1 1 HS2 or 2H1 1 S22. 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). Anode:

H2 S-S0 1 2H1 1 2e2 E0 5 2 0:141 V

ð9:12Þ

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 (S2O322), sulfite (SO322), and sulfate (SO422). From an economical point of view, the preferred pathway is the conversion to elemental sulfur (2e2) in contrast to sulfate (8e2). 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 Ag21, Ce41, Fe21, and (VO2)1 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. H2 S 1 OH2 -H2 O 1 HS2 Anode: Cathode:

9.4.2

HS2 1 OH2 -S 1 H2 O 1 2e2 2H2 O 1 2e2 -H2 1 2OH2

ð9:13Þ ð9:14Þ ð9:15Þ

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

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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

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.

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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). 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 passed 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 passed 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 (Gerc¸el 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

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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 terms 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 6 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 LaxSr12xVO32δ (LSV) anodes for H2S-containing 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/LSM-YSZ 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

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must certainly be extended in order for H2S electro-oxidation to be applied at an industrial scale.

9.5 9.5.1

Biogas upgrading approach and its challenges CO2 electroreduction

The economic feasibility of various CO2R products needs to be taken into account 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 e2. 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 S´anchez 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

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TABLE 9.2 Advantages and challenges of various designs of electrolyzers used for CO2R. Reactor configuration

Advantages

Batch reactors

G

Onecompartment cells

G

Simple design Convenient for small-scale experiments

Drawbacks G G G

G

Twocompartment cells (H-cell)

G

G

G

Continuousflow reactors

Membrane reactor

G

G

G

G

G

Microfluidic reactor

G

G

G

Convenient to study half-cell reactions Convenient for small-scale experiments Separated compartments avoid product crossover and reoxidation

G

Relatively easy to upscale Different configurations possible Pressurization relatively easy Existing know-how, due to similarities with water electrolyzers Straightforward stack construction

G

No membrane included Fast screening of catalysts under various operating conditions Better control of flooding/dry-out phenomena at electrodes

G

G G

G

G

G

G

Poor mass transfer Poor mixing Reoxidation of reaction products at the anode Contamination of working electrode more likely Poor mass transfer Poor mixing Additional losses due to membrane resistance

Little knowledge on the effects of various types of membranes Membrane may lead to additional resistance or product crossover Membrane cost and stability issues

Less effective separation of reaction products from electrodes Difficult to pressurize compared with the membrane reactor Scale-up more difficult compared with the membrane reactor (Continued )

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TABLE 9.2 (Continued) Reactor configuration Solid-oxide electrolyzer

Advantages G

G

G

Decrease in overpotentials Increase in charge transfer Long-term CO2 electrolysis feasible (B500 h)

Drawbacks G

G

G

Extreme conditions required Impurities in the gas stream may lead to electrode passivation Techno economic feasibility for CO2 reduction unknown

Source: Reprinted after S´anchez, 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.

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. 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 support 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 (S´anchez 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

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inhibition and both electrochemical and mechanical failures were observed. Modifying anodes with CeO2 resulted in only a slight improvement in durability. A real-biogas-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 well-established 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

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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 (HS2) and CO2 (HCO32). 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% 6 1% to 100% 6 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 h/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. 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).

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.

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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 co-electrolysis of CO2 and steam accompanied by pretreatment steps was the most preferred method in terms 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 ch/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 e-gas (Germany). 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 value-added 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

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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 (S´anchez 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 supported 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.

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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

Chapter outline 10.1 Introduction 255 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 10.4 Comparison of the methods for reducing the content of volatile organic silicon compounds in biogas 10.5 Conclusions and future perspective References

281

281 286 288

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). 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

Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00010-6 © 2021 Elsevier Inc. All rights reserved.

255

TABLE 10.1 Nomenclature and basic properties of VOSCs detected in biogas. Chemical name

Abbreviation

Chemical formula

Vapor pressure, kPa, at 25 C

Water solubility, mg/dm3 at 25 C

Boiling point,  C, at 101.3 kPa

LogKOC, at 25 C

LogKOA, at 25 C

Hexamethylcyclotrisiloxane

D3

Si3-O3-(CH3)6

1.16

1.571

135

3.55

4.056

Octamethylcyclotetrasiloxane

D4

Si4-O4-(CH3)8

0.14

0.055

176

4.17

4.28

Decamethylcyclopentasiloxane

D5

Si5-O5-(CH3)10

0.03

0.046

210

4.60

4.95

Dodecamethylcyclohexasiloxane

D6

Si6-O6-(CH3)12

0.003

0.005

245

5.08

5.77

Hexamethyldisiloxane

L2

Si2-O-(CH3)6

5.613

2.881

101

3.75

2.89

Octamethyltrisiloxane

L3

Si3-O2-(CH3)8

0.445

0.15

152

4.37

3.75

Decamethyltetrasiloxane

L4

Si4-O3-(CH3)10

0.05

0.007

195

4.80

4.66

Dodecamethylpentasiloxane

L5

Si5-O4-(CH3)12

0.013

0.0003

232

5.54

4.49

Trimethylsilanol

TMS-OH

Si-(CH3)3-OH

1.84

42,560

99

1.24

3.62

Tetramethylsilane

TMS

Si-(CH3)4

95.03

215

23 28

1.69

1.00

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

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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

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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, supporting 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

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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 B10% 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 230 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

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concentrations of VOSCs (Bseveral hundred of mg/m3) or in the case of simultaneous CO2 removal (CO2 boiling point at 0.1 MPa is 278.5 C), or the production of liquefied biomethane (CH4 boiling point at 0.1 MPa is 2161.5 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 (Cabrera-Codony 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. The tests confirmed the clear effect of the Brunauer-Emmett-Teller (BET) surface area and the volume of micropores (1 2 nm) and small

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TABLE 10.2 Physical parameters of the selected granular ACs used for VMSs removal from biogas. Parameter

Number of ACs tested

Value

Unit

Source

Specific surface area (BET)

6

650 1370

m2/g

Oshita et al. (2010)

12

850 1757

Cabrera-Codony et al. (2014)

5a

909 1250

Cabrera-Codony et al. (2018)

5

b

11

1082 2190 262 1573

Yu et al. (2013)

c

3

302 ; 879; 1104

Nam et al. (2013)

12

819 1323

Vagenknechtov´a 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

1

0.6 3.0

Total pore volume

6

0.35 0.94

nm

Oshita et al. (2010) Nam et al. (2013) Sigot et al. (2016)

cm3/g c

Oshita et al. (2010)

3

0.45; 0.47; 0.18

Nam et al. (2013)

12

0.39 1.38

Vagenknechtov´a et al. (2017)

1

0.67

Ortega and Subrenat (2009)

12

0.45 1.52

Cabrera-Codony et al. (2014)

5a

0.42 1.10

b

0.91 1.58

Cabrera-Codony et al. (2018)

5

AC, activated carbon; VMSs, volatile methylsiloxane. a Steam activation. b H3PO4 activation. c 2.5% KOH impregnation.

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

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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), ˚ molecular sieve CH4 adsorption into AC is two times greater than into a 5 A 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 B5 C) and compressed (B2.5 MPa) biogas, showed that AC under these conditions

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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. 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

TABLE 10.3 Results of VMS adsorption capacity tests for various granulated ACs. Kind of tested biogas

Kind of adsorbate

Adsorption capacity, g/kg

Source

Real sewage biogas

Total VMSs

8.5 21.5

Hepburn et al. (2015)

5

de Arespacochaga et al. (2014)

10 15

Wheless and Pierce (2004)

D4 1 D5

63

Vagenknechtov´a et al. (2017)

Synthetic biogas (CH4/ CO2 5 50/50, vol/vol)

D4

43 334

Synthetic biogas (carrier gas: air)

L2

Synthetic biogas (carrier gas: N2)

22

D4

D5

a

225

Gong et al. (2015) b

Ortega and Subrenat (2009)

10 100

Gislon et al. (2013)

53

Sigot et al. (2016)

56 192

Matsui and Imamura (2010)

50 404

Oshita et al. (2010)

50 531 106

D6

18

Total VMSs

155 307

L2

47 123

D4

36 90

D5

47 169

AC, activated carbon; VMSs, volatile methylsiloxane. a BET: 432 m2/g, micropore volume (Vmic): 0.18 cm3/g. b BET: 1573 m2/g, Vmic 5 0.62 cm3/g.

Yu et al. (2013)

110 350

Kajolinna et al. (2015)

Nam et al. (2013)

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PART | II Physiochemical upgrading systems

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 (, 230 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.

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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 self-ignition 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%

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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 B5 nm. TABLE 10.4 Physical parameters of the selected granular SGs used for VMS removal from biogas. Parameter Specific surface area (BET)

Average pore diameter

Value 690

Unit 2

m /g

Sigot et al. (2016)

607

Ortega and Subrenat (2009)

470

Jung et al. (2017)

717

Oshita et al. (2010)

656

Nam et al. (2013)

350

Montanari et al. (2010)

2.2

nm

2.34 Total pore volume

Source

0.24 0.36

SG, silica gel; VMS, volatile methylsiloxane.

Soreanu et al. (2011) Nam et al. (2013)

3

cm /g

Ortega and Subrenat (2009) Nam et al. (2013)

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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

TABLE 10.5 The results of VMS adsorption capacity tests for various SGs. Kind of tested biogas

RH, %

Kind of adsorbate

Adsorption capacity, g/kg

Source

Synthetic biogas, carrier gas: N2

0

D4

216 259

Sigot et al. (2016)

70

21

ND

104

Matsui and Imamura (2010)

B100

Jung et al. (2017)

0

D5

25

77

50

B42

100

35

,10

L2, D5

.100

Schweigkofler and Niessner (2001)

ND

Total VMSs

202

Nam et al. (2013)

L2

17

D4

56

D5

129 91

Kajolinna et al. (2015)

D6

20

Synthetic biogas, carrier gas: air

70

L2

B130

Ortega and Subrenat (2009)

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

ND

D3

230

Montanari et al. (2010)

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

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PART | II Physiochemical upgrading systems

may therefore be a better adsorbent than AC for landfill gas treatment, which usually contains more L2 and less H2S than sewage gas. 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

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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). 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). Adsorption capacity and regeneration of zeolites The D4 adsorption process using popular 13X molecular sieves has been studied, for example, in Germany (Schweigkofler and Niessner, 2001), Japan (Matsui and Imamura, 2010), France (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 Cabrera-Codony 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. 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.

TABLE 10.6 Classification and parameters of the selected zeolite adsorbents. Parameter

Kind of zeolite Synthetic

Natural

3A (KA)

4A (NaA)

5A (CaA)

10X (CaX)

13X (NaX)

ZSM5

Modernite (Na)

Chabazite (Na, Ca)

Clinoptilolite (NaX)

Silicalite

Nominal pore opening, nm

0.3

0.4

0.5

1.0

1.0

0.55

0.4

0.49

0.35

0.53

Si/Al

0.7 1.2

5

4.5 5.0

1.6 3.0

4.2 5.2

.100

1.0 1.5

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TABLE 10.7 Physical parameters of the selected ZEs used for VMS removal from biogas. Parameter Specific surface area (BET)

Average pore diameter Total pore volume

Type

Value

Unit 2

m /g

Source

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

Oshita et al. (2010)

Six synthetic ZEs

370 910

Cabrera-Codony et al. (2017)

ND

3 9

Hydrophobic ZE

3

DAY40

0.33

ZSM-5

0.21

UCT-15

0.21

Hydrophobic ZE

0.51

nm

Ortega and Subrenat (2009)

Soreanu et al. (2011) Oshita et al. (2010)

cm3/g

Ortega and Subrenat (2009) Jiang et al. (2016)

Oshita et al. (2010)

ZEs, zeolites; VMS, volatile methylsiloxane.

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 B400 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

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TABLE 10.8 The results of VMS adsorption capacity tests for the various ZEs. Kind of tested biogas

Kind of ZE

Kind of adsorbate

Adsorption capacity, g/kg

Source

Synthetic biogas, carrier gas: N2

13X

D4

113

Sigot et al. (2016)

77

Matsui and Imamura (2010)

8A

4

13X (MS544)

D3

276

Montanari et al. (2010)

Hydrophobic ZE

D4

51

D5

52

Oshita et al. (2010)

19

Jung et al. (2017)

D4

41

Jiang et al. (2016)

ZE (cbv720) ZSM-5

Synthetic biogas, carrier gas: air

UCT-15

77

Clinoptilolite

11

Six synthetic ZEs

28 143

DAY40

L2

140

Cabrera-Codony et al. (2017)

Ortega and Subrenat (2009)

ZEs, zeolites; VMS, volatile methylsiloxane.

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). 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 (Al120-8h) 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

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TABLE 10.9 Physical parameters of the selected AAs used for VMS removal from biogas. Parameter Specific surface area (BET)

Average pore diameter

Total pore volume

Type

Value

Commercial AA

201

Synthetic Mesoporous AA (Al120-8h)

314

Commercial AA—Daejung (Korea)

250

Commercial AA

3.4

Synthetic Mesoporous AA (Al120-8h)

4.3

Commercial AA—Daejung (Korea)

6.7

Commercial AA

0.26

Synthetic Mesoporous AA (Al120-8h)

0.46

Commercial AA—Daejung (Korea)

0.38

Unit 2

m /g

Source Zhong et al. (2017) Nam et al. (2013)

nm

Zhong et al. (2017) Nam et al. (2013)

cm3/g

Zhong et al. (2017) Nam et al. (2013)

AA, activated Al2O3; VMS, volatile methylsiloxane.

TABLE 10.10 The results of VMS adsorption capacity tests for the various AAs. Kind of tested medium

Kind of AA

Kind of adsorbate

Adsorption capacity, g/kg

Source

Synthetic biogas, carrier gas: N2

Commercial AA

D4

127

Zhong et al. (2017)

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

Commercial AA— Daejung (Korea)

Synthetic Mesoporous AA (Al120-8h)

AA, activated Al2O3; VMS, volatile methylsiloxane.

168

L2

8

D4

34

D5

104

Total VMS

146

Nam et al. (2013)

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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.

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. TABLE 10.11 Physical parameters of the selected PAs used for VMS removal from biogas. Parameter Specific surface area (BET)

Average pore diameter

Total pore volume

Type

Value

Unit 2

RS1

936

PDVB

831

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)

9

XAD-4 (PS/DVB)

4

XAD-7 (acrylic ester)

9

PDVB

1.5 60

PDVB

1.7

PDVB-VI

1.2 1.8

PAs, polymer adsorbents; VMS, volatile methylsiloxane.

m /g

Source Oshita et al. (2010) Jafari et al. (2015)

nm

Jafari et al. (2015) 3

cm /g

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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. 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 (B1000 g/kg).

TABLE 10.12 The results of VMS adsorption capacity tests for the various PAs. Kind of tested medium

Kind of AP

Kind of adsorbate

Adsorption capacity, g/kg

Source

Synthetic biogas, carrier gas: N2

RS1

D5

500

D4

300

Oshita et al. (2010)

PDVB-VI-0.12

2370a, 2360b a

b

PDVB

1951 , 1940

PDVB-VI-0.24

1586a, 1582b

PDVB-VI-0.50

1384a, 1381b

P(DVB-ACAM)

2220

VMS, volatile methylsiloxane; PAs, polymer adsorbents a RH 5 0%. b RH 5 50%.

Jafari et al. (2015)

Noshadi et al. (2016)

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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 contact 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 La¨ntela¨ et al. (2012)] can reduce VMSs by more than 99%. The absorption of VMSs in hydrocarbon oils was tested in the United Kingdom and Germany (along with halides removal) but without promising results [according to EA (2010), B60% 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

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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 [B6.5 h/kg—according to Cormos et al. (2018)]. The Kryosol process is also described (EA, 2010), and uses chilled (B 2 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 B61% and B82%, respectively. The cost of oil, however, is quite significant (14 h/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.

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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 (B7% 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

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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 contact 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.

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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 contact 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 admixtures 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.

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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 terms 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 support 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 Summary review of the methods and techniques for removing VOSCs from biogas. Technology

Removal efficiency and performance

Advantages

Drawbacks

Cost considerations

1

2

3

4

5

Pretreatment methods

,90%

Increasing the intensity of biogas production, improving the sludge dewatering

They can pose a threat to air quality

Low or moderate costs can be expected

Condensation methods

,50%

Simple technology. Extension of an adsorbent’s or absorbent’s lifetime at the next stage of biogas treatment

Low efficiency in relation to VOSCs. They can only be the initial stage of biogas purification

Low investment costs, medium operational costs

Freezing methods—up to 230 C

,90%

Easy operation. High removal efficiency. Simultaneous removal of VOSCs with other impurities (H2S, NH3, halides)

Freezing of apparatus—periodic defrosting is required. High energy consumption

High investment and operating costs. It can be profitable only for high concentrations of VOSCs

Adsorption into activated carbon (AC)

,99% Real biogas: 5 63 g/kg. Synthetic biogas (N2): 10 404 g/kg

Simple, mastered technology—easy operation, widely available and relatively cheap adsorbent, possibility of simultaneous removal of other undesirable biogas components

Nonselectivity. Competing adsorption, mutual displacement of VMSs and their polymerization into AC surface, which blocks AC pores and limits its regeneration. Risk of self-ignition during regeneration. The need for gas pretreatment. Greater CH4 adsorption capacity than in the case of SG and 5A molecular sieve

Relatively low costs, for example, for 190 m3/h of biogas stream: investment costs: 38,000 h; operational costs: 4500 h/y (de Arespacochaga et al., 2014) or 0.003 0.023 h/kWh (Ajhar et al., 2010). Predominant part of operating costs is the spent AC replacement. The price of commercial ACs: 2 3 h/kg

(Continued )

TABLE 10.13 (Continued) Technology

Removal efficiency and performance

Advantages

Drawbacks

Cost considerations

1

2

3

4

5

Adsorption into conventional silica gel (SG)

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

Features of the SG: mechanical strength, high thermal resistance, nonflammability, chemical, and biological inertness. Greater selectivity and better regeneration susceptibility than AC

High affinity for water and the need for deep biogas drying. SG is not suitable as an H2S adsorbent

Relatively low costs. The system is cheaper than with AC, provided that the biogas is dried , 10% RH. The price of commercial SGs: 1 2 h/kg

Adsorption into zeolites (ZEs)

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

High affinity for H2S. Higher hydrophobicity and thermal resistance compared to AC and SG. Better adsorbent for simultaneous removal of H2S and VOSCs than AC and SG

Pore sizes are usually smaller than the size of the VOSC molecules

Relatively low costs. Sample price of zeolite based on halloysite is around 1 h/kg

Adsorption into commercial alumina (AA)

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

AA contains mainly mesopores with a diameter of 3 7 nm, optimal for VMSs adsorption (excl. L2). High thermal resistance. Relatively easy regeneration, without the noticeable effect of VMS polymerization

Relatively small specific surface area

Relatively low costs

Adsorption into polymer adsorbents (PA)

,99% Synthetic biogas (N2): 500 g D5/kg, 300 2370 g D4/kg

High VMS adsorption capacity, chemically homogeneous structure, relatively large specific surface area and pore volume, usually high hydrophobicity, ease of regeneration

Low thermal resistance

High production costs—the most expensive VOSC adsorbent

(Continued )

TABLE 10.13 (Continued) Technology

Removal efficiency and performance

Advantages

Drawbacks

Cost considerations

1

2

3

4

5



(at 100 C), lack of VMS polymerization Absorption in water

,50%

Side effects of processes such as condensation drying, wet dedusting, absorption of CO2, NH3, and H2S in aqueous solutions

Due to the poor solubility of most VOSCs, it is only suitable for TMSOH, TMS, L2, and D3 removal

Negligible costs when VOSC removal occurs while other processes are running

Absorption in hydrocarbons, alcohols, and hydrocarbon oils

,95%

Effective removal of VOSCs together with other VOCs and halides. Easy regeneration

Ignition risk and toxicological hazards. Risk of desorption and solvent evaporation due to an increase in biogas temperature or flow rate. The need for cooling

High investment costs, high costs of absorbents, expensive regeneration

Absorption in Selexol

,95%

Advantages of the absorbent: low vapor pressure, weak binding with VOSCs, low affinity for CH4, low viscosity, chemical stability, low freezing point, noncorrosivity and nontoxicity, easy regenerability. Can be used for simultaneous biogas drying, absorption of H2S and CO2

Risk of desorption due to an increase in biogas temperature or flow rate. Stripped gas during regeneration of the absorbent should be captured and neutralized

Relatively expensive (B 6.5 h/kg)

Absorption in silicone oil

,99% for D4, B 60% for L2, B 80% for L3

Advantages of the absorbent: as above and high thermal stability.

As above

Cost of oil: B14 h/kg

(Continued )

TABLE 10.13 (Continued) Technology

Removal efficiency and performance

Advantages

Drawbacks

Cost considerations

1

2

3

4

5

Can be used for simultaneous removal of halides and other VOCs Membrane techniques

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

High selectivity, possibility of continuous operation, modularity, and compactness of the plant. It allows simultaneous drying of biogas and removal of VMSs. This technique is beneficial for biogas enrichment, before introducing into natural gas grid. Does not cause sewage or solid waste. The process can be intensified by combining with absorption in one apparatus

High energy demand to power the compressors or vacuum pumps. The need for preliminary removal of aerosols, oil vapors, and chemical impurities that can block and damage the membrane, risk of CH4 losses. The permeate requires further processing

Moderate operational costs. Investment costs are relatively low

Biological methods using biotrickling filter (BTF)

In relation to D4: 43% at a residence time of 19.5 min and 74% at a residence time of 13.2 min

These methods are expected to be environmentally friendly. It is possible to intensify absorption by adding a nonbiodegradable and nonmiscible organic solvent, and intensify biodegradation by adding biosurfactants or specific enzymes. Possibility of simultaneous removal of H2S and VOSCs in one device

A large mass exchange surface and long contact time are required. These methods were tested only in laboratory scale, using synthetic biogas. Low process flexibility, long time needed for adaptation of microorganisms

Low operational costs can be expected

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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 terms 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

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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 contact 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.

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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.

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Jung, H., Lee, D. Y., Jurng, J., 2017. Low-temperature regeneration of novel polymeric adsorbent on decamethylcyclopentasiloxane (D5) removal for cost-effective purification of biogases from siloxane. Renew. Energy 111, 718 723. Kajolinna, T., Aakko-Saksa, P., Roine, J., Ka˚ll, L., 2015. Efficiency testing of three biogas siloxane removal systems in the presence of D5, D6, limonene and toluene. Fuel Process. Technol. 139, 242 247. Kazuyuki, O., Masaki, T., Tadao, M., Hiroshi, K., Nobuo, T., Akira, K., 2007. Behavior of siloxanes in a municipal sewage-treatment plant. J. Jpn Sew. Work. Assoc. 44 (531), 125 138. La¨ntela¨, J., Rasi, S., Lehfinen, J., Rintala, J., 2012. Landfill gas upgrading with pilot scale water scrubber: performance assessment with absorption water recycling. Appl. Energy 92, 307 314. Lee, S.-H., Cho, W., Song, T.-Y., Kim, H., Lee, W.-J., Lee, Y.-C., et al., 2001. Removal process for octamethylcyclotetrasiloxane from biogas in sewage treatment plant. J. Ind. Eng. Chem. 7 (5), 276 280. Li, Y., Zhang, W., Xu, J., 2014. Siloxanes removal from biogas by a lab-scale biotrickling filter inoculated with Pseudomonas aeruginosa S240. J. Hazard. Mater. 275, 175 184. Liu, Y.-H., Meng, Z.-Y., Wang, J.-Y., Dong, Y.-F., Ma, Z.-C., 2019. Removal of siloxanes from biogas using acetylated silica gel as adsorbent. Pet. Sci. 16, 920 928. Matsui, T., Imamura, S., 2010. Removal of siloxane from digestion gas of sewage sludge. Bioresour. Technol. 101 (1), S29 S32. Miguel, G.S., Lambert, S.D., Graham, J.D., 2002. Thermal regeneration of granular activated carbons using inert atmospheric conditions. Environ. Technol. 23 (12), 1337 1346. Mito-oka, Y., Horike, S., Nishitani, Y., Masumori, T., Inukai, M., Hijikata, Y., et al., 2013. Siloxane D4 capture by hydrophobic microporous materials. J. Mater. Chem. A. 1, 7885 7888. Montanari, T., Finocchio, E., Bozzano, I., Garuti, G., Giordano, A., Pistarino, C., et al., 2010. Purification of landfill biogases from siloxanes by adsorption: a study of silica and 13X zeolite adsorbents on hexamethylcyclotrisiloxane separation. Chem. Eng. J. 165, 859 863. Nam, S., Namkoong, W., Kang, J., Park, J., Lee, N., 2013. Adsorption characteristics of siloxanes in landfill gas by the adsorption equilibrium test. Waste Manage 33 (10), 2091 2098. Noshadi, I., Kanjilal, B., Jafari, T., Moharreri, E., Khakpash, N., Jiang, T., et al., 2016. Hydrophobic mesoporous adsorbent based on cyclic amine divinylbenzene copolymer for highly efficient siloxane removal. RSC Adv. 6, 77310 77320. Ortega, D.R., Subrenat, A., 2009. Siloxane treatment by adsorption into porous materials. Environ. Technol. 30 (10), 1073 1083. Oshita, K., Ishihara, Y., Takaoka, M., Takeda, N., Matsumoto, T., Morisawa, S., et al., 2010. Behaviour and adsorptive removal of siloxanes in sewage sludge biogas. Water Sci. Technol. 61 (8), 2003 2012. Oshita, K., Omori, K., Takaoka, M., Mizuno, T., 2014. Removal of siloxanes in sewage sludge by thermal treatment with gas stripping. Energ. Conv. Manage. 81, 290 297. Oshita, K., Fujime, M., Takaoka, M., Fujimori, T., Appels, L., Dewil, R., 2015. Siloxane removal and sludge disintegration using thermo-alkaline treatments with air stripping prior to anaerobic sludge digestion. Energ. Conv. Manage. 96, 384 391. Petersson, A., Wellinger, A., 2009. Biogas upgrading technologies developments and innovations. Task 37 - Energy from biogas and landfill gas. IEA Bioenergy. Popat, S.C., Deshusses, M.A., 2008. Biological removal of siloxanes from landfill and digester gases: opportunities and challenges. Environ. Sci. Technol. 42 (22), 8510 8515. 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.

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Chapter 11

Technologies for removal of hydrogen sulfide (H2S) from biogas Anish Ghimire1, Raju Gyawali2, Piet N.L. Lens3 and Sunil Prasad Lohani4 1

Department of Environmental Science and Engineering, Kathmandu University, Dhulikhel, Nepal, 2Nepal Electricity Authority, Government of Nepal, Kathmandu, Nepal, 3UNESCO—IHE Institute for Water Education, Delft, The Netherlands, 4Department of Mechanical Engineering, Kathmandu University, Dhulikhel, Nepal

Chapter outline 11.1 Introduction 11.2 Technologies for removal of biogas contaminants 11.3 Physicochemical removal technologies 11.3.1 Absorption process 11.3.2 Adsorption process 11.3.3 Membrane separation 11.4 Ex situ removal using sulfur-oxidizing microorganisms 11.4.1 Biological air filtration

295 297 298 298 300 304

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11.4.2 Microalgal removal of H2S 11.5 In situ H2S removal 11.5.1 In situ microaeration 11.5.2 Dosing iron salts/oxides into the digester 11.6 Combined chemical-biological processes 11.7 Comparison of H2S removal techniques 11.8 Conclusions References

311 312 312 313 314 314 316 317

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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00011-8 © 2021 Elsevier Inc. All rights reserved.

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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 (Gutie´rrez-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. H2 S"H1 1 HS2 "2H1 1 S22

ð11:1Þ

H2 S 1 Fe-FeS 1 H2

ð11:2Þ

FeS 1 2H1 -Fe21 1 H2 S

ð11:3Þ

In addition, H2S reduces the Fe

31

present in rust,

2Fe31 1 H2 S-2Fe21 1 S 1 2H1

ð11:4Þ

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.

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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

Ex situ H2S removal

(A)

Biological treatment

Removal using sulfide-oxidizing microorganisms

Removal using microalgae

Physicochemical treatment

Chemical absorption

Physical adsorption

Pressure swing adsorption

Membrane separation

Water scrubbing

Adsorption on activated carbon

Adsorption on iron oxides/hydroxides Fe2O3/(Fe(OH)3)

(B)

In situ H2S removal

Iron chloride (FeCl2) dosing into the digester (in situ H 2S precipitation)

Microaeration (dosing air/oxygen into the digester)

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

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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.

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) (Thra¨n et al., 2014). Due to its simplicity and effectiveness, physicochemical absorption is the most common technology for simultaneous H2S and CO2 removal (Thra¨n 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-pass scrubbing and regenerative absorption.

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FIGURE 11.2 Schematic of water scrubbing system for H2S removal.

In a single-pass 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

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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

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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 HS2 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 terms of pH, its value should be higher than 5 to ensure the effectiveness of the H2S removal (the right concentration of HS2 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 ZnO-impregnated 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 H 2S 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

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PART | III Biological upgrading systems

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: FeO 1 H2 S-FeS 1 H2 O

ð11:5Þ

Fe2 O3 1 3H2 S-Fe2 S3 1 3H2 O

ð11:6Þ

1 FeS 1 O2 -FeO 1 S 2

ð11:7Þ

3 Fe2 S3 1 O2 -Fe2 O3 1 3S 2

ð11:8Þ

Regeneration:

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: G

G

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 H 2S 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).

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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 pass 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). 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

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

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PART | III Biological upgrading systems

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 pass 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

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

single-stage and two-stage membrane filtration processes. Losses of 5%
10% methane are possible as the CH4 may also pass 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 terms of CH4 purity and recovery compared to multistaged systems (CH4 purity .99% and recovery .95%).

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 - HS2 1 H1;

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PART | III Biological upgrading systems

3. H1 produced in step (2) is consumed by CO322 that is diffusing toward the high-pressure side by the reaction H1 1 CO322- HCO32; 4. HS2 and HCO32 diffuse from the high-pressure to the low-pressure side of the film; 5. HS2 combines with H1 near the low-pressure side of the film by the reaction HS2 1 H1 - H2S; 6. H1 for step (5) is supplied by HCO32 by the reaction HCO32- H1 1 CO322; 7. CO322 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 support 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 pass 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-Guzm´an et al., 2016). Biofilters (BFs), bioscrubbers (BSs), and biotrickling filters (BTFs) are proven cost-effective and environmentally friendly biological air filtration technologies (Dumont, 2015; Barbusi´nski 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). BFs consist of a packed bed which is generally composed of peat, compost, and synthetic media with high nutrient availability, high buffer

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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).

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 (NO32) 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.

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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 (SO-NR) microorganisms such as Thiobacillus denitrificans, which converts H2S to elemental sulfur and sulfate (depending on the nitrate/sulfide ratio) (Eq. 11.9) using nitrate (NO32) as the electron acceptor (Soreanu et al., 2008b). H2 S 1 CO2 1 Nutrients 1 Nitrate - Biomass 1 Sulfur and=or Sulfate

ð11:9Þ

1 Water 1 Nitrite and=or Nitrogen gas 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): 1 2 1 HS2 1 1:23NO2 3 1 0:093NH4 1 0:438HCO3 1 0:573H 1 0:027CO2

-0:093C5 H7 O2 N 1 0:614N2 1 SO22 4 1 0:866H2 O ð11:10Þ 1 2 1 HS2 1 0:350NO2 3 1 0:013NH4 1 0:059HCO3 1 1:4H 1 0:004CO2

-0:013C5 H7 O2 N 1 0:175N2 1 S0 1 1:21H2 O ð11:11Þ 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-NO32 consumed/g H2S removed). Complete denitrification vs. complete H2S oxidation (ratio N/S 5 1.6) 22 1 5H2 S 1 8NO2 3 -5SO4 1 4N2 1 4H2 O 1 2H

ð11:12Þ

Complete denitrification vs. partial H2S oxidation (ratio N/S 5 0.4) 0 2 5H2 S 1 2NO2 3 -5S 1 N2 1 4H2 O 1 2OH

ð11:13Þ

Partial denitrification vs. complete H2S oxidation (ratio N/S 5 4) 22 2 1 H2 S 1 4NO2 3 -SO4 1 4NO2 1 2H

ð11:14Þ

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Partial denitrification vs. partial H2S oxidation (ratio N/S 5 1) 0 2 H2 S 1 NO2 3 -S 1 NO2 1 H2 O

ð11:15Þ

Overall equation: 0 22 2 2 1 15NO2 3 1 12H2 S-9H2 O 1 6S 1 6SO4 1 5NO2 1 5N2 1 2OH 1 4H ð11:16Þ

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 S0 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 S0 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; Fern´andez 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: H2 S 1 0:5O2 -S0 1 H2 O 1 H2 S 1 2O2 -SO22 4 1 2H



ΔG 5 2 169:35 kJ=mol

ð11:17Þ

ΔG 5 2 732:6kJ=mol

ð11:18Þ

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 (Toma`s et al., 2009). The SO422/S0 ratio produced varies considerably depending on the supplied O2/H2S ratio. With the

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increasing H2S loading rates the sulfate concentration decreases, thus lowering the SO422/S0 ratio. The major drawback of elemental sulfur formation is the clogging of the reactor bed resulting in a pressure drop (Toma`s et al., 2009). Toma`s et al. (2009) evaluated the technoeconomic feasibility of a fullscale 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 contact 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 (h3.2 per kg H2S treated) over chemical treatment (h5.8 per kg H2S treated) of biogas for H2S removal with a cost saving of about h2.6 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, France), 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 twostep process is summarized in Eqs. 11.19 and 11.20: H2 SðgÞ 1 NaOHðaqÞ-NaHSðaqÞ 1 H2 O

ð11:19Þ

NaHSðaqÞ 1 1=2O2 -SðsÞ 1 NaOHðaqÞ

ð11:20Þ

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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 sulfur-oxidizing 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; Mun˜oz 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 (Mun˜oz 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 2 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.

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The microalgal H2S removal technology is yet to be demonstrated in fullscale 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): 2H2 S 1 O2 -2S 1 2H2 O

ð11:21Þ

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 (Jen´ıcˇ ek 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 Biogas

Biogas Aeration

Liquid recirculation

Aeration

Biogas recirculation

Effluent/ digestate Digesting mixture

Influent/ feed

Effluent/ digestate Digesting mixture

Influent/ feed

A

B

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.

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adopted for H2S removal from biogas (Khoshnevisan et al., 2017; Mun˜oz et al., 2015). 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. Jen´ıcˇ ek 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 yellow-whitish 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 Fe21 or Fe31). 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. Fe21 1 H2 S-FeSk 1 2H1

ð11:22Þ

2FeðOHÞ3 1 3H2 S-Fe2 S3 k 1 6H2 O

ð11:23Þ

2FeCl3 1 H2 S-Sk 1 2FeCl2 1 2HCl

ð11:24Þ

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

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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). First stage : H2 S 1 Fe2 ðSO4 Þ3 -S0 1 2FeSO4 1 H2 SO4

ð11:25Þ

Secondary stage : 2H2 SO4 1 4FeSO4 1 O2 -2Fe2 ðSO4 Þ3 1 2H2 O

ð11:26Þ

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% Fe31 formation. In addition, the optimal glucose concentration in the nutrient medium was 0.1% for the optimal growth of T. ferrooxidans CP9. Pagella and De Faveri (2000) studied the combined action of a chemical-biological process for the removal of H 2S 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

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TABLE 11.1 Advantages and disadvantages of techniques for removal of H2S (Awe et al., 2017; Ryckebosch et al., 2011; Angelidaki et al., 2018; Sun et al., 2015; Mun˜oz et al., 2015). H2S removal process Biological

Absorption

Advantages and features

Disadvantages

In situ microaeration

No separate reactors needed for H2S removal, treatment cost of 0.0037 (pure oxygen) and h0.0026 m3 (air) of biogas treated (Khoshnevisan et al., 2017)

May lead to deposition of elemental sulfur on the reactor system, control of air/ O2 dose or O2/H2S is challenging, needs additional upgrading steps, requires careful designing of reactor headspace

Biological air filtration (BF, BS, BTF)

Operation at ambient temperature and pressure, low operation and maintenance cost, high removal possible ( . 97% CH4), possible to bring H2S to ,50 ppm, cost associated with BTF at h0.013 to 0.016 m3 of biogas treated (Khoshnevisan et al., 2017)

Controlling O2/N2 in biogas implies difficult and additional upgrading steps in the case of aerobic technologies, elemental sulfur deposition on the media requires periodic cleaning, and also strict control of process parameters like pH, moisture, and nutrient conditions, concentration in the treated biogas could be still high (100
300 cm3/m3)

Microalgal technologies

H2S and CO2 could be simultaneously removed, operation at ambient temperature and pressure, use of algae biomass for energy generation

No/limited full-scale experiences, O2/N2 in biogas needs additional upgrading steps

Water scrubbing

Proven and simple technology, economical if water is available, CO2 can be removed along with H2S, energy requirements of 0.3
0.9 kWh/Nm3 clean biogas

Requires high pressure and low temperature, expensive, lower flexibility toward variation of input gas, clogging due to bacterial growth

Organic solvents (amine)

High efficiency ( . 99% CH4), low CH4 losses (,0.1%)

High (CAPEX and O&M) costs, heat requirements for regeneration, foaming problems

Absorption with NaOH and FeCl3

Low electricity requirements, low CH4 losses, operating cost around h0.03/m3 of biogas

High (CAPEX and O&M) costs, not regenerative (standalone process)

(Continued )

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TABLE 11.1 (Continued) H2S removal process

Advantages and features

Disadvantages

Adsorption

Activated carbon

High efficiency (H2S , 3 cm3/m3), high purification rate, low operation temperature, high loading capacity, cost of treatment around h3.85/kg H2S

Regeneration requires high temperature, (450 C) and residues are present up to 850 C, requires pretreatment for O2 and H2O, CH4 losses

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

Simple technology, low investment cost, high removal efficiency ( . 99% CH4) along with removal of mercaptans

Sensitive for water (pretreatment required), risk of ignition of the impregnated wood chips (regeneration is exothermic)

PSA

Compact technique, tolerant to impurities, efficient technology (96
98% CH4), energy requirements of 0.25
1 kWh/Nm3 clean biogas, flexibility for small applications

High (CAPEX and O&M) costs, CH4 losses (,4%), extensive process control

Gas
gas; gas
liquid

Compact process, removal of more CO2 and H2, energy requirements of 0.14
0.26 kWh/Nm3 clean biogas, less than 0.6% CH4 loss, 96
98% CH4 recovery

High investment cost, expensive operation and maintenance, biogas requires pretreatment, less operational experience of gas
liquid membranes

Membrane

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

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).

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 terms of performance, cost effectiveness, flexibility, and complexity of operation. Therefore the selection of technology requires careful examination of the required gas quality to avoid

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unnecessary treatment costs. Physicochemical processes are much more efficient in terms 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 support the sustainability of biogas utilization.

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32. Available from: https://doi.org/ 10.11648/j.ijrse.s.2015040101.14. Rasi, S., Veijanen, A., Rintala, J., 2007. Trace compounds of biogas from different biogas production plants. Energy. 32, 1375
1380. Available from: https://doi.org/10.1016/j. energy.2006.10.018. Rezakazemi, M., Heydari, I., Zhang, Z., 2017. Hybrid systems: combining membrane and absorption technologies leads to more efficient acid gases (CO2 and H2S) removal from natural gas. J. CO2 Util. 18, 362
369. Available from: https://doi.org/10.1016/j.jcou.2017.02.006. Ryckebosch, E., Drouillon, M., Vervaeren, H., 2011. Techniques for transformation of biogas to biomethane. Biomass Bioenergy 35, 1633
1645. Available from: https://doi.org/10.1016/J. BIOMBIOE.2011.02.033. Salas, B., Wiener, M.S., Badilla, G.L., Beltran, M.C., Zlatev, R., Stoycheva, M., et al., 2012. H2S pollution and its effect on corrosion of electronic components. In: Badilla, M., Valdez, G.L., Schorr, B. (Eds.), Air Quality
New Perspective. INTECHOPEN, pp. 263
286. Available from: https://doi.org/10.5772/39247. San Miguel, G., Lambert, S.D., Graham, N.J.D., 2001. The regeneration of field-spent granularactivated carbons. Water Res. 35, 2740
2748. Shareefdeen, Z., Herner, B., Singh, A., 2005. Biotechnology for air pollution control—an overview. Biotechnol. Odor Air Pollut. Control. Springer, pp. 3
15. Soreanu, G., Be´land, M., Falletta, P., Edmonson, K., Seto, P., 2008a. 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. 7, 543
552. Available from: https://doi. org/10.1139/S08-023. Soreanu, G., Be´land, M., Falletta, P., Edmonson, K., Seto, P., 2008b. Laboratory pilot scale study for H2S removal from biogas in an anoxic biotrickling filter. Water Sci. Technol. 57, 201
207. Available from: https://doi.org/10.2166/wst.2008.023. Sun, Q., Li, H., Yan, J., Liu, L., Yu, Z., Yu, X., 2015. Selection of appropriate biogas upgrading technology—a review of biogas cleaning, upgrading and utilisation. Renew. Sustain. Energy Rev. 51, 521
532. Available from: https://doi.org/10.1016/j.rser.2015.06.029. Syed M., Soreanu G., Falletta P., Be´land M., Removal of hydrogen sulfide from gas streams using biological processes—a review, (2006). Thra¨n, D., Billig E., Persson T., Svensson M., Daniel-Gromke J., Ponitka J., et al., BiomethaneStatus and Factors Affecting Market Development and Trade, 2014. http://task40.ieabioenergy.com/wpcontent/uploads/2013/09/t40-t37-biomethane-2014.pdf. Toma`s, M., Fortuny, M., Lao, C., Gabriel, D., Lafuente, J., Gamisans, X., 2009. Technical and economical study of a full-scale biotrickling filter for H2S removal from biogas. Water Pract. Technol. 4. Available from: https://doi.org/10.2166/wpt.2009.026. Wellinger A., Lindberg A., Biogas Upgrading and Utilization, 2007. http://www.iea-biogas.net/ files/daten-redaktion/download/publi-task37/Biogasupgrading.pdf. Zheng, Y., Brown, B., Neˇsi´c, S., 2014. Electrochemical study and modeling of H2S corrosion of mild steel. Corrosion. 70, 351
365. Available from: https://doi.org/10.5006/0937. Zicari S.M., Removal of hydrogen sulfide from biogas using cow - manure compost, Cornell University, 2003. Zulkefli, N.N., Masdar, M.S., Jahim, J., Majlan, E.H., 2016. Overview of H2S removal technologies from biogas production. Int. J. Appl. Eng. Res. 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

Chapter outline 12.1 Biogas upgrading 12.2 Hydrogen generation and utilization 12.3 Methanation 12.4 Microbial basis for biomethanation 12.4.1 Methanogens 12.4.2 Processes in anaerobic digestion 12.5 Reactor configurations 12.5.1 In situ biomethanation 12.5.2 Ex situ biomethanation 12.6 Factors controlling biomethanation 12.6.1 Mass transfer of H2

321 324 326 327 327 328 330 331 332 337 337

12.6.2 12.6.3 12.6.4 12.6.5

Temperature Growth requirements pH and CO2 Bacterial interaction and competition 12.7 Reactor design for biological methanation 12.7.1 Continuous stirred tank reactor 12.7.2 Trickle-bed reactors 12.8 Future perspectives and applications 12.9 Conclusions Abbreviations list References

343 344 346 348 350 350 352 354 356 357 357

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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00012-X © 2021 Elsevier Inc. All rights reserved.

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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. 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

TABLE 12.1 Standards for natural gas quality (Sun et al., 2015). EU

US

Min. CH4 concentration

70
98 % mol

93.5
95.5 % mol

Max. CO2 concentration

1
8 % mol

2
3 % mol 0.2
3 % mol

Max. O2 concentration

0.01
1 % mol

Max. N2 concentration

2
10 % mol

Max. H2 concentration

0.1 % mol

0.1 % mol 3

Max. H2S concentration

2
15 mg/m

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

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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.

used for valorization of waste gas CO2 by capture and conversion to CH4 through the reaction with H2 [Eq. (12.1)] (Go¨tz 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). 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: 4H2 1 CO2 -CH4 1 2H2 O

ð12:1Þ

The reaction stoichiometry reveals that significant reducing equivalents are needed to reduce the one-carbon atom from its most oxidized form (CO2; oxidation state 14) to its most reduced form (CH4; oxidation state 24). 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-togas technologies enable energy production and consumption to be balanced (IRENA, 2019a).

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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 steam-reforming of fossil CH4 [Eq. (12.2)]. CH4 1 2H2 O-4H2 1 CO2

ð12:2Þ

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. 2H2 O-2H2 1 O2

ð12:3Þ

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 (Go¨tz 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 supported 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 panels, 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 supporting 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

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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..

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, methanation-based CH4 produced from H2 [Eq. (12.1)] offers a versatile option for immediate large-scale production, application, and storage of an energyrich molecule. CH4 is hence directly applicable as both a transport fuel and for large-scale 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 5 241.82 kJ/mol) versus that of CH4 (ΔHcΘ 5 802.62 kJ/mol), it is clear that the conversion of four moles of H2 to one mole of CH4 results in an energy loss of B17% (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 (Go¨tz 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

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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.

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 (B300 C) and high pressure (5220 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 (Go¨tz 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

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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.

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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 (Leve´n 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). 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

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

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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. 2HCO3 2 1 4H2 1 H1 -CH3 COO2 1 4H2 O

ð12:4Þ

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 H2-producing 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. 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; Stru¨bing et al., 2017). The use of this complex inoculum for biomethanation has often been

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TABLE 12.2 Gibbs free energy of acetogenic processes under standard conditions and at low H2 concentration (Madigan et al., 2019a). Free energy change Reactions

Free energy change ΔG0’ (kJ/reaction)a

ΔG (kJ/reaction) PH2 5 1024 atm

Glucose 1 4H2O - 2 acetate2 1 2HCO32 1 4H1 1 4H2

2207

2319

Butyrate2 1 2H2O - 2 acetate2 1 H1 1 2H2

148.2

217.6

Propionate2 1 3H2O - acetate2 1 HCO32 1 H1 1 H2

176.2

25.5

2 Ethanol 1 2H2O - 2 acetate2 1 2H1 1 4H2

119.4

237.0

a

Standard conditions, solutes 1 M gases, 1 atm., 25 C.

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 terms 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

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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.

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.

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

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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 (Garcia-Robledo 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 lowcost 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

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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 mixedculture 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; Porte´ 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). In general, ex situ biomethanation expands the possibilities to control the reactions compared to the in situ configuration. However, most of the factors

TABLE 12.3 In situ and ex situ biomethanation reactor performances based on continuous operation. CH4 production rates are given at standard conditions (1 bar, 273.15K). In cases where studies did not specify the physical basis for gas volumes, these were assumed to be listed at process conditions. Yields are estimated based on data provided. CCH4, methane concentration in product gas; YCH4/H2, CH4 yield from H2 based on stoichiometric conversion (4:1) (Modified from Jensen et al., 2020b). In situ/ ex situ

Methane production rate (NL/L/d)

CCH4 (%)

YCH4/H2 (%)

Reactor configuration

Working volume (L)

Culture

In situ

0.74

96

.91

Continuous stirred tank reactors (CSTR)

0.6

In situ

0.45

78

N.A.

CSTR

In situ

0.38

65

92

In situ

0.30

85

Ex situ

690.8

Ex situ

Pressure [bar(g)]

Temperature ( C)

CO2 source

References

Mixed

55

Biogas

Luo and Angelidaki (2013)

4.5

Mixed

55

Biogas

Zhu et al. (2019)

CSTR

3.5

Mixed

55

Biogas

Luo et al. (2012)

95

CSTR

2.0

Mixed

55

Biogas

Bassani et al. (2015)

23

N.A.

CSTR

1.0

Pure

65

CO2

Nishimura et al. (1992)

233.6

96

96

CSTR

1.5

Pure

65

CO2

Peillex et al. (1990)

Ex situ

171.4a

95

.95

CSTR

7000a

Pure

63

CO2

Electrochaea GmbH (2017); Sveinbjo¨rnsson and Mu¨nster (2017)

Ex situ

144.0

23

96

Trickle bed

0.05

Pure

65

CO2

Jee et al. (1987)

2

8

Ex situ

137.6

85

Ex situ

135.8

25

.97

CSTR

1.0

Pure

Ex situ

122.7

58

100

Trickle bed

0.11

Pure

Ex situ

a

72.0

95

.95

CSTR

7000

Ex situ

61.0

15

100

CSTR

1.0

Pure

Ex situ

53.8

7

N.A.

CSTR

3.0

Pure

Ex situ

41.1

14

N.A.

Trickle bed

0.10

Ex situ

33.7

44

100

Trickle bed

Ex situ

26.4

90

98

Ex situ

15.4

99

Ex situ

11.0

Ex situ

9.8

N.A.

CSTR

10

Pure

CO2

Seifert et al. (2014)

60

CO2

Schill et al. (1996)

65

CO2

Jee et al. (1988a)

63

Biogas

Electrochaea GmbH (2017); Sveinbjo¨rnsson and Mu¨nster (2017)

65

CO2

Rittmann et al. (2012)

60

CO2

Martin et al. (2013)

Pure

65

CO2

Jee et al. (1988b)

0.23

Mixed

35

Synthetic biogas

Dupnock and Deshusses (2017)

Trickle bed

0.75

Mixed

37

CO2

Savvas et al. (2017)

99

Trickle bed

58.1

Mixed

55

CO2

Stru¨bing et al. (2017)

97

99

Trickle bed

0.70

Mixed

37

CO2

Dupnock and Deshusses (2019)

61

N.A.

CSTR

3.0

Pure

60

Biogas

Martin et al. (2013)

a

Pure

1

8

0.2

65

(Continued )

TABLE 12.3 (Continued) In situ/ ex situ

Methane production rate (NL/L/d)

CCH4 (%)

YCH4/H2 (%)

Reactor configuration

Working volume (L)

Culture

Ex situ

4.0

91

92

CSTR

0.6

Mixed

Ex situ

3.0

98

100

CSTR

5000

Mixed

Ex situ

2.3

83

76

Trickle bed

0.29

Mixed

Ex situ

1.6

97

N.A.

Trickle bed

5.8

Mixed

Ex situ

1.5

95

99

Trickle bed

61

Ex situ

1.5

95

99

Trickle bed

Ex situ

1.4

87

96

Ex situ

1.2

99

Ex situ

0.7

Ex situ

0.4

a

Estimated from available data.

Pressure [bar(g)]

Temperature ( C)

CO2 source

References

55

Synthetic biogas

Luo and Angelidaki (2012)

60

Biogas

IEA (2018)

52

Synthetic biogas

Sieborg et al. (2020)

37

Biogas

Rachbauer et al. (2016)

Mixed

37

CO2

Burkhardt et al. (2015)

1.0

Mixed

54

Synthetic biogas

Porte´ et al. (2019)

Trickle bed

4.0

Mixed

54

CO2

Alitalo et al. (2015)

93

Fixed bed

7.8

Mixed

35

CO2

Lee et al. (2012)

96

92

Fixed bed

0.85

Mixed

55

Synthetic biogas

Bassani et al. (2017)

79

N.A.

CSTR

4.5

Mixed

60

CO2

Pokorna et al. (2019)

4
14

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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). 

Cl 5

Pg H

ð12:5Þ

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. 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.

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TABLE 12.4 Henry’s law constant for solubility in water at 298.15K (mol/kg/bar) for different gases relevant to biomethanation processes (NIST, 2020). 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 3 1025

Yes

However, the acid
base properties of CO2 mean that its overall dissolution capacity is pH-dependent 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)]. CO2 1 H2 O$H2 CO3 $HCO3 2 1 H1 1 $CO3 22 1 2H1

ð12:6Þ

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.

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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.

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 liquidside mass transfer coefficient for diffusion of the solubilized gas through the stagnant liquid zone.

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.

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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 rate-limiting process in biomethanation. 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)]. J 5 2DH2 ;H2 O 3

dCH2 dx

ð12:7Þ

where J is the absolute molar flux of H2 (mol/s/m2), DH2 ;H2 O is the diffusion coefficient of H2 in H2O (m2/s), dCH2 =dx 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

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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 J 0 5 kg 3 ðP
Pi Þ

ð12:8Þ

J 0 5 kl 3 ðCi
CÞ

ð12:9Þ

Liquid film flux

0

2

J is the molar flux of gas (mol/m /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 terms 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: k5

DH2 x

ð12:10Þ

where k denotes kg or kl (m/s), DH2 is the diffusion coefficient of H2 in gas (DH2 ;gas ) or liquid (DH2 ;H2 O ), 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)]. 1 1 1 5 1 Kl H 3 kg kl

ð12:11Þ

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: 

rH2 5 kl 3 a 3 ðCl 2 Cl Þ

ð12:12Þ

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PART | III Biological upgrading systems

where rH2 is the H2 gas
liquid mass transfer rate (mol/m3/s), kl is the liquid side mass transfer coefficient used as an approximation for the overall mass transfer coefficient (Kl ) (m/s), a is the volumetric gas
liquid interfacial area  (m2/m3), Cl is the concentration of dissolved H2 in equilibrium with its gas phase partial pressure [Eq. (12.5)] (mol/m3), and Cl is the actual concentration of dissolved H2 in the liquid phase. Several operational factors affect the individual terms 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 account. Actual methanogenic activity in the boundary layer would effectively decrease the distance a H2

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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

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PART | III Biological upgrading systems

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)]. 0:131CO2 1 0:004HCO3 2 1 0:004NH4 1 1 0:5H2 - 0:115CH4 1 0:004C5 H7 O2 N 1 0:266H2 O

ð12:13Þ

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.

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TABLE 12.5 Elemental composition of cellular biomass from methanogens (Scherer et al., 1983). The values are derived from the analysis of 10 different methanogenic pure cultures. 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

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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/HCO32/CO322 [Eq. (12.6)], NH41/NH3, and CH3COOH/CH3COO2, and only to a minor degree Mg21, Ca21, 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

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and the metabolic water consequently contributes more to the total liquid content as a consequence (Stru¨bing et al., 2017). The buffer of primary importance for controlling pH in biomethanation systems is the pool of DIC represented by CO2/HCO32/CO322, 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 HCO32 during common pH conditions of biomethanation reactors (pH 7
8), so CO2 depletion is accompanied by H1 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 (Stru¨bing 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 (Stru¨bing et al., 2017). The decrease in pH observed by Stru¨bing 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.

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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 supported by Jensen et al. (2019) who found that up to 63 6 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 (ΔG0 5 2131.0 kJ/mol) compared to homoacetogenesis (ΔG0 5 294.9 kJ/mol) (Schink, 1997). On an enzyme level, the substrate affinities also indicate that

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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 B8 Pa, Methanosarcina barkeri 150 Pa, M. marburgensis B8 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 (Trˇcek 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)].  2 ½A  pH 5 pKa 1 log ð12:14Þ ½HA where pKa is the pH value at which the protonated acid (HA; acetic acid) and unprotonated acid (A2; 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.

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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 diffuser which together provide and disperse H2 bubbles in the liquid broth. The diffuser 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 6 3% to 68 6 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 diffuser, from a ceramic diffuser with pore size of 0.5
1.0 mm to a column diffuser with a pore size of 14
40 μm resulted in increased CH4 content in the product gas from 68% to 75% due to

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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 B7 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, Germany, 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):  b P a k L a 5 K 3 Vs 3 ð12:15Þ V 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

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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 CSTR-based biomethanation reactors are associated with a significant power input for stirring (Savvas et al., 2017; Stru¨bing 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 (Delhome´nie 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 tricklebed 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).

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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 terms 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 contact 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 mixedculture 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 (Delhome´nie 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

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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). 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 panels—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 supports 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

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TABLE 12.6 Strengths, weaknesses, opportunities, and threats analysis of biomethanation (Modified from Agneessens, 2018). Strengths G

G

G G G

G

Biomethane is a flexible and mature energy carrier giving it versatile and immediate application Use of existing gas grid infrastructure enables cheap, large-scale, and longterm energy storage Use of existing biogas infrastructure Valorization of waste CO2 streams Valorize renewable power outside of the electricity sector, which only covers a minor part of the total energy demand Reduced dependency of biomass for production of renewable carbonbased fuel

Opportunities G

G

G

G G

Binding and more ambitious targets for renewable transport fuels Use of biomethane as a building block for production of base chemicals through intermediates such as synthetic gas Breakthrough in electrolysis investment costs Increasing taxes on CO2 emission Continued implementation of renewable power production based on wind and sun

Weaknesses G

G

G

G

Low H2 solubility entails high energy demand (gas
liquid mixing) and/or demand for construction of additional reactor volume (ex situ step) Additional investment cost compared to direct usage of electricity or H2 as an energy carrier Inevitable energy loss compared to direct usage of H2 gas as an energy carrier Requires continued implementation of renewable power production based on wind and sun

Threats G

G

G

G

G

Continued low fossil natural gas prices or decreased value of green gas certificates No significant breakthrough in gas
liquid mass transfer technologies Lack of political awareness and unadapted legislation Continued high electricity prices caused by transmission and system operation levies Development of competing power-toX technologies

transport possible (IRENA, 2018a). Being a power-to-gas technology, biomethanation-based biogas upgrading is hence supporting 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

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involved in the biomethanation process. Go¨tz 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%; Go¨tz 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 supported by political action to promote the attractiveness of substituting fossil natural gas for renewable biomethane, similar to existing support 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 supported 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

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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 DIC CSTR SWOT

volatile fatty acid dissolved inorganic carbon continuous stirred tank reactor strengths, weaknesses, opportunities, threats

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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 1

Department of Biological and Chemical Engineering, Aarhus University, Aarhus, Denmark, Separation and Conversion Technology, Flemish Institute for Technological Research (VITO), Mol, Belgium 2

Chapter 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 13.4 Economical insights 13.5 Prospective and challenges 13.6 Conclusion References

370 375 376 378 378

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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00013-1 © 2021 Elsevier Inc. All rights reserved.

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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 2 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. The processes follow the reaction schemes presented below: Cathodic reduction CO2 1 8H1 1 8e2 -CH4 1 2H2 O

ð13:1Þ

CO2 1 4H2 -CH4 1 2H2 O

ð13:2Þ

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365

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.”

Anodic oxidation 2H2 O-O2 1 4H1 1 4e2

ð13:3Þ

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,

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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 (Schro¨der 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. CE 5

F3C3V 3z Ð to t Idt

ð13:4Þ

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 (t-to) 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

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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 ΔGR 3 n EE 5 Ð to t I 3 Udt

ð13:5Þ

where ΔGR is the theoretical thermodynamic Gibbs free energy for the reduction reactions (Eqs. 13.1
13.3) and n is the number of moles of CO2 that are reduced during the reaction time (t 2 to). ΔGR 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 U is the voltage applied to the electrochemical system. The product between I and U 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 U 5 ER 1 EA 1 RTot 3 I 5 ER 1 EA 1 ðRct;A 1 Rct;C 1 Rmt 1 ROhm Þ 3 I ð13:6Þ

where ER 5 ΔGR =ðzFÞ is the theoretical minimum electric potential needed for driving the reduction reaction (Eqs. 13.1
13.3). EA is the activation overpotential needed for initiating the reactions and is typically associated with the electrode materials and their catalytic characteristics (Eq. 13.1). RTot 3 I is the extra total overpotential for actually driving the reaction, while RTot is the total electrical resistance that can be divided into several contributions. ROhm is related to ohmic losses associated with electrical resistance in the electrodes or resistance of the conduction of ions through the ion-exchange membrane. Rmt 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 Rct;A and Rct;C are the charge transfer resistances, related to the anode and cathode, respectively, and describe losses in the electrochemical reactions. ER ; EA ; Rct;A ; Rct;C , and ROhm can to a good approximation be assumed to be independent of the current, while the mass transfer/concentration polarization term (Rmt ) has a strong dependence on the electrical current, in particular when the mass transfer limit is approached. In the case of BES, where

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typical current densities (electrical current per unit electrode area) are below 20 mA/cm2 and in many cases also includes stirring/convection, Rmt 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 (ROhm ), electrolyte and membrane (ROhm ), electrodes and catalyst (EA ; Rct;C ), microbes (EA ; Rct;C ), and stirring/convection (Rmt ). An overall increase in the EE is therefore a matter of identifying the terms 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 members of Methanosaeta, which has shown the capability to accept electrons directly for the CO2 reduction to methane when tested in BES (Blasco-Go´mez 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 contact 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).

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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 hydrogen-mediated hydrogenotrophic methanogenesis, while upgrading biogas (Batlle-Vilanova 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-Go´mez 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). 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

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.

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PART | III Biological upgrading systems

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 Recent studies on bioelectrochemical systems employed for the conversion of biogas CO2 at laboratory scale. Electrode material

Operating mode

Reactor design

Inoculum source

Current density/ draw

Upgrading/ improvement

CE

Reference

Graphite

Batch in situ

H cell

1 A/m2

CO2 removal .10%

.82.6

Xu et al. (2014)

Batch ex situ

Biogas from AD fed to H cell

Anaerobic granular sludge from UASB reactor treating synthetic brewery waste water

0.4 A/m2

CO2 removal .10%

.85.3

Continuous in situ

Single-chamber

3 A/m2

CO2 removal .8%

80
90

Stainless steel of reactor acts as electrode

Batch in situ

Coupled MEC-AD (single-chamber barrel shaped made from stainless steel reactor)

From MFC

Ng

98% CH4

66.

Bo et al. (2014)

Graphite block

Batch in situ

2 chamber BES

Effluent from AD

0.07 A/m2

65%
85% CH4

75.3

0.2 A/m2

Three fold faster in CH4 production rate

69.9

BatlleVilanova et al. (2015)

200 A/m2

Significant amount of CO2 was removed as acetate

99

Continuous in situ

MWCNTRVC

Batch

H cell

Enriched mix culture obtained from WWTPs

Jourdin et al. (2016a)

(Continued )

TABLE 13.1 (Continued) Electrode material

Operating mode

Reactor design

Inoculum source

Current density/ draw

Upgrading/ improvement

CE

Reference

Stainless steel of reactor as electrode

Batch in situ

Barrel-shaped singlechamber BES

Geobacter-containing inoculum from MEC reactor and Methanosarcina sp.

304.3 A/m3

24% CH4 compared to control

74.6

Yin et al. (2016)

RVC

Batch in situ

Single-chamber BES

Digested sludge from WWTP

0.112 A/m2

8%
9% CH4 yield

41.6

Gajaraj et al. (2017)

Carbon felt

Batch in situ

Two-chamber BES


0.1 A

98% CH4

72.8

Batch in situ

Single-chamber BES

Thermophilic anaerobic sludge obtained from MSW treated digester


0.002 A

56% CH4

80.9

Liu et al. (2017)

Titanium woven wire mesh coated with Pta

Batch in situ

MESC

Sludge from domestic WWTP

1.7 A/m2

97.5% CH4

81%

Jin et al. (2017)

Graphite granular

Batch in situ

Three-compartment BES

Activated sludge enriched in hydrogenotrophic chemoautotrophic microorganisms from WWTP

0.087 A

90% CO2 removal

86.4%

Zeppilli et al. (2017)

Cu-Ni- and Fe-coated graphite carbon

Batch in situ

AD-MEC singlechamber reactor

FWTP

Ng

Four times faster than AD

Ng

Park et al. (2018)

Stainless steel woven mesha

Batch in situ

MESC

Sludge from domestic WWTP

1.49 A/m2

99
100% CH4

Ng

Kokkoli et al. (2018)

Graphite granular

Batch in situ

Two-sided cathode compartment in between anode compartment

Sludge from thermophilic AD

0.15 A

63% CO2 removal

125 6 7

0.13 A

49% CO2 removal

108 6 15

Zeppilli et al. (2019)

Graphite granulara

Batch in situ

2 chamber BES

Granular sludge from UASB

0.2 A/m3

0.23m3/m3 day CH4 production

33 6 10

Cerrillo et al. (2017)

Carbon fiber brush

Batch in situ

Single-chamber BES

Sludge from WWTP

0.018 A

Two fold higher CH4 production than control

Ng

Lee et al. (2019)

Graphite plate

Continuous in situ

Continuous flow MEC reactor

Anaerobic sludge from MWWTP

Ng

.90% CH4

97.6

BatlleVilanova et al. (2019)

Graphite plate

Continuous in situ

H cell

Sludge from WWTP

Ng

.90% CH4

97.6

Liu et al. (2019)

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. a Maximum value calculated.

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PART | III Biological upgrading systems

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-ofconcept 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 dual-chamber 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 NH41, OH2, HCO32, and CH3COO2 (Zeppilli et al., 2017, 2019). Another study applied a four-compartment BES system with an additional twochamber 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.

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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 h/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 early-stage 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 h/kWh CH4), compared to the chemical catalysis methanation systems (0.130
0.159 h/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 h/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 h. 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 h/m3 or 0.1 h/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.

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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 LCH4/L/d, using a doublechambered 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 h/cm2 the stack/electrode cost alone would exceed 34 Mh. 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-Go´mez 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

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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 contact 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 lab-scale 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 h/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-Go´mez et al., 2017; Christodoulou et al., 2017). Thus, product diversification of

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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-Go´mez 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 supported 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.

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Yin, Q., Zhu, X., Zhan, G., Bo, T., Yang, Y., Tao, Y., et al., 2016. Enhanced methane production in an AD and microbial electrolysis cell coupled system with co-cultivation of Geobacter and Methanosarcina. J. Environ. Sci. (China) 42, 210
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Chapter 14

Photosynthetic biogas upgrading: an attractive biological technology for biogas upgrading Vijay Kumar Garlapati1, Swati Sharma1 and Surajbhan Sevda2 1

Department of Biotechnology and Bioinformatics, Jaypee University of Information Technology, Waknaghat, India, 2Department of Biotechnology, National Institute of Technology Warangal, Warangal, India

Chapter outline 14.1 Introduction 383 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 14.6.1 Light intensity 14.6.2 Media pH 14.6.3 Temperature 14.6.4 Biogas composition 14.6.5 Gas flow rate 14.7 The future prospects 14.8 Conclusion References

392 392 394 394 395 395 396 401 401

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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00014-3 © 2021 Elsevier Inc. All rights reserved.

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[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; Mun˜oz 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

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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 France, 3% in Austria, 4% in Sweden, and 6% in Switzerland and Germany) (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 cryogenicbased 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; Mun˜oz 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

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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/SO422, 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 photosynthetic-bacterial 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)] (Mun˜oz et al., 2015). CO2 1 Sunlight ðphotonsÞ 1 nutrients-O2 1 CH1:63 N0:14 O0:43 P0:006 S0:005 1 Waste heat

ð14:1Þ

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

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(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 (Mun˜oz 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; Gonz´alez-Camejo 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] (Lo´pez et al., 2018). H2 SðgÞ ’-H2 SðliqÞ

ð14:2Þ

H2 SðliqÞ 12O2 -SO4 2- 12H1

ð14:3Þ

H2 SðliqÞ 1 0:5O2 -S0 k 1 H2 O

ð14:4Þ

Aerobic oxidation

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

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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.

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.

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

TABLE 14.1 Photosynthetic-based biogas upgradation studies with concomitant waste water treatment. TN, total nitrogen; TP, total phosphorus. Biogas composition

Microalgae/ microalgae
cocultivation

Upgraded CH4 content

Treated waste water results (% removal of constituents)

References




Chlorella vulgaris cultivated with fungi

87.03%
88.01%

92.17% (COD), 89.83% (TN), 90.31% (TP)

Zhao et al. (2019)

Raw biogas (CH4—61.23%, CO2—34.69%)

C. vulgaris cocultivated with activated sludge

83.24%

79.86% (COD), 80.23% (TN), 89.37% (TP)

Gao et al. (2018)

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

Microalgae and bacteria

76%
99%




del Rosario Rodero et al. (2018)

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

C. vulgaris cocultivated with Ganoderma lucidum

86%

83% (COD), 81% (TN), 84% (TP)

Guo et al. (2018)

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

Picochlorum sp. cultivated with Halospirulina sp.

.96%

52%
55% (TN), 12%
29% (TP)

Franco-Morgado et al. (2018)

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

C. vulagris with activated sludge

83%

81% (COD), 83% (TN), 83% (TP)

Guo et al. (2018)

Synthetic biogas with various concentrations of carbon dioxide

C. vulgaris cocultivated with Ganoderma lucidum

93%

86.08% (COD), 85.69% (TN), 86.17% (TP)

Wang et al. (2017a,b)

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

Chlorella minutissima

96%

85% (COD), 85% (TN), 100% (TP)

Toledo-Cervantes et al. (2017a,b)

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

C. vulgaris cocultivated with Ganoderma lucidum

92%

70% (COD), 76% (TN), 78% (TP)

Cao et al. (2017)

Crude biogas (CH4—64.58% CO2—31.72% H2S—,0.005%)

C. vulgaris cocultivated with Ganoderma lucidum

90%

74% (COD), 76% (TN), 77% (TP)

Zhang et al. (2017)

Crude biogas (CH4—64.58% CO2—31.72% H2S—,0.005%)

C. vulgaris cocultivated with activated sludge

93%

72% (COD), 75% (TN), 74% (TP)

Zhang et al. (2017) (Continued )

TABLE 14.1 (Continued) Biogas composition

Microalgae/ microalgae
cocultivation

Raw biogas (CH4—57.32% CO2—39.25%)

C. vulgaris cocultivated with fungi

Crude biogas (CH4—63.84% CO2—31.02%)

Chlorella sp.

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

Upgraded CH4 content

Treated waste water results (% removal of constituents)

References

79.74% (COD), 74.28% (TN), 85.37% (TP)

Guo et al. (2017a, b)

92.87%

92.67% (COD), 80.87% (TN), 79.33% (TP)

Yan and Zheng (2013)

Scendesmus sp.

70%

83% (TN), 78% (TP)

Prandini et al. (2016)

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

Chlorella sp.




85.23% (COD), 87.10% (TN), 92.40% (TP)

Yan and Zheng (2013)

Crude biogas (CH4—67.32% CO2—34.45% H2S—, 0.005%)

C. vulgaris

80.04%

59.27% (COD), 51.32% (TN), 63.22% (TP)

Zhao et al. (2015)

Crude biogas (CH4—67.32% CO2—34.45% H2S—, 0.005%)

Nannochloropsis Oleoabundans

80.06%

61.03% (COD), 51.74% (TN), 54.33% (TP)

Zhao et al. (2015)

Raw biogas (CH4—61.75% CO2—35.28%)

Scendesmus obliquus

94.41%

93.08% (COD), 84.12% (TN), 86.76% (TP)

Ouyang et al. (2015)

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

C. vulgaris

98.9%

51% (COD), 37% (TN), 86% (TP)

Serejo et al. (2015)

Crude biogas (CH4—67.32% CO2—34.45% H2S—, 0.005%)

Scendesmus sp.

82.64%

63.12% (COD), 53.04% (TN), 59.14% (TP)

Zhao et al. (2015)

Raw biogas (CH4—58.67% CO2—37.54%)

Scendesmus obliquus

88.25%

61.58
75.29% (COD), 58.39
74.63% (TN), 70.09
88.79% (TP)

Xu et al. (2015)

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

Chlorella sp.

92.74

85.35% (COD), 77.98% (TN), 73.03% (TP)

Zhao et al. (2013)

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391

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.

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. The flat-plate PBRs are robust, high surface-to-volume ratio configurations with easy operation and superior efficiency. The low reactor thickness of platetype 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).

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PART | III Biological upgrading systems

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; Mun˜oz 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 value-added 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). µ 5 µmax :I=ðKE 1 IÞ

ð14:5Þ

Photosynthetic biogas upgrading Chapter | 14

393

where µmax denotes the maximum specific growth rate and KE is the light half-saturation 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 redlight 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).

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PART | III Biological upgrading systems

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-Arde`vol 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-Arde`vol 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 (FrancoMorgado 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) (Mar´ın 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

Photosynthetic biogas upgrading Chapter | 14

395

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; Gonz´alez-S´anchez 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). µ5

µmax S Ks 1 S

ð14:6Þ

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 (Gonz´alez-S´anchez 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).

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PART | III Biological upgrading systems

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 terms 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 , L/ G , 67) had a proven optimum for CO2 (80%) and H2S (100%) removal (Serejo et al., 2015). Another biogas upgradation study utilizing an HRAP and algal-bacterial system reported 0.5 L/G ratio and cocurrent flow operation resulting in better upgradation results (Toledo-Cervantes et al., 2017a,b). In another study by Rodero et al., 2019b, a step-wise increase in flow rate results in enhanced biogas upgradation results. The optimal results were seen with gas flow rates of 60
120 mL/min, utilizing the 15 C
35 C, and 100
1500 mg/L inorganic carbon sources. The recent research findings of photosynthetic biogas upgradation with different reactor systems and utilized cultivation parameters and obtained biomass productivities have been summarized in Table 14.2.

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

TABLE 14.2 Photosynthetic biogas upgrading studies using different algal systems/photobioreactors. HRT, hydraulic retention time. Algal reactor configurations

Algal cultivation parameters

RE-CO2

CH4(%)/O2 (%)/N2 (%)

Biomass productivity

References

16.8 L glass photobioreactor filled with slurry and addition of different concentrations of strigolactone GR24 Biogas composition—59.87% CH4, 35.96% CO2, 0.45% O2, 3.71% H2O

Light intensity 200 µmol/m2/s/ 12 h:12 h light and dark Nitrogen source—urea NaH2PO4  2H2O, KH2PO4 as phosphorus source

64.05% with 1027 M GR24




0.097 6 0.008 g/ L/d

Shen et al. (2020)

Tubular photobioreactor Synthetic biogas composition 70% CH4, 29.5% CO2, 0.5% H2S.

Photosynthetically active radiance883 µmol/m2/s/12 h:12 h light and dark periods Synthetic flue gas flow rate—20 mL/min Liquid gas ratio—0.5

96%

97.2% CH4

15.8 6 4.3 g/m2/d

A´ngeles et al. (2020)

Open 50 L photobioreactor Connected with bubble column operated continuous lyInjected biogas—65 6 1.5 % CH4; 32.0 6 1.9 % CO2

Light intensity of 25, 50, 75, and 100 µmol/m2/s/12 h:12 h light and dark, L/G ratio—0.6

89%
93%

Less than 1% O2

0.06 g/L/d

Meier et al. (2017)

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

Light intensity 1500 6 600 µmol/ m2/s/14 h:10 h light and dark flow rate 1.6 m3/m2/h liquid to biogas flow rate ratio—0.3
1.0 biogas sparged into AC 40 mL/min

98.8%

96.2 % CH4, 0.7% O2, 2.7% N2

15 g/m2/d

ToledoCervantes et al. (2017a,b) (Continued )

TABLE 14.2 (Continued) Algal reactor configurations

Algal cultivation parameters

RE-CO2

CH4(%)/O2 (%)/N2 (%)

Biomass productivity

References

Outdoor high rated algal pond attached with bubble absorption column Biogas composition—CO2 29.5%, H2S 0.5% and CH4 70%

Biogas flow rate— 74.9 L/d centrate flow rate—3.2 L/d liquid gas ratio-0.5

50%
95%

94% CH4, 0.7
1.2% O2

15 g/m2/d

Posadas et al. (2017)

25 L high rated algal pond connected with absorption bubble column Synthetic biogas 30% CO2, 0.5% H2S

Photosynthetically active radiance—500 µmol/m2/s Hydraulic retention time—9.5 days

91.5%




0.0193 g/L/d

FrancoMorgado et al. (2017)

Two interconnected cylindrical glass incorporated with biogas and slurry with 60
70% methane content

Light intensity of 200 µmol/m2/s/ 16 h:8 h light and dark periods

66.93%
88.27%

64.58% CH4, 0.54 % O2

0.171 g/L/d

Zhang et al. (2017)

Two connected cylindrical glass have synthetic biogas with varying CO2 (25.27%
55.17%) concentration and slurry

Light intensity 200 µmol/m2/s/ 14 h:10 h light and dark periods

81.39% Cocultivation with fungi in mixed light




0.424 g/L/d

Wang et al. (2017a,b)

Closed glass Photo bioreactor of 30 cm height and 20 cm diameter used Biogas composition—65%
73% CH4, 20%
25% CO2, 1800
3100 ppmv H2S

Red light emission diode 148.5 µmol/m2/s/12 h:12 h light and dark

66.7%

64.7% CH4, 17.8% O2

1.1 6 0.2 g/L

Prandini et al. (2016)

Photo bioreactor with 20 L crude gas and 4 L slurry H2S— ,50ppm

Mixed LED as light with Photosynthetic photon flux density 200 µmol/m2/s/12 h:12 h light and dark

62%

79%
83% CH4, 0.62% O2




Zhao et al. (2015)

75 L photobioreactor connect with bubble column filled with real biogas 72% CH4, 28% CO2

Light intensity 100 µmol/m2/s 4.75 L/d feed flow, gas flow rate— 100 ml/min

93%

1.2% O2

0.03 g/L/d

Meier et al. (2015)

180 L HRAP connected with bubble column Synthetic biogas— 30% CO2, 69.5% CH4, 0.5% H2S

Light intensity of 104 µmol/m2/s/ 16 h:8 h light and dark periods Synthetic gas mixture sparged in to AC—0.2
1.0 m3/m2/h, 7.4 days HRT

86%

88% CH4,1% O2, 6% N2




Serejo et al. (2015)

Interconnected 180 L raceway photobioreactor with bubble column filled with synthetic biogas—30% CO2, 69.5% N2 and 0.5% H2S

Gro-lux fluorescent lamp of 80 µE/ m2/s Gas flow rate—38 m3d

86%

0.2% O2




Bahr et al. (2014)

Raceway plant with recirculation rate 0.22 m/s filled with 10.6% CO2 flue gas

Gas flow rate—100 L/min

96%

.15% O2

0.088 g/L/d

De Godos et al. (2014)

Bubble column photobioreactor Biogas composition—69% CH4, 20% CO2 and 0.005% H2S

Light intensity—300 µmol/m2/s Gas flow rate—0.05—0.3 vvm

74%
85%

86%
91% CH4

0.3
0.32 g/L/d

Kao et al. (2012a)

420 L raceway connect with bubble column reactor

Gas flow rate—0.93 L/min

82%
83%







Putt et al. (2011) (Continued )

TABLE 14.2 (Continued) Algal reactor configurations

Algal cultivation parameters

RE-CO2

CH4(%)/O2 (%)/N2 (%)

Biomass productivity

References

Bubble column reactor filled with real biogas 38%
80% CH4, 19%
62% CO2, and 0.2% H2S

Light intensity—1000 µE/m2/s Gas flow rate—0.5
10 L/min Ammonia as nitrogen for microalgae cultivation

74%
85%

.3.5% O2

2.6
3.8 g/L/d

Douˇskov´a et al. (2010)

Tubular photobioreactor Biogas composition—41% CO2, 57.5% CH4 and 0.05% H2S

Light intensity 35
100 µmol/m2/s

98%

50%
53% CH4 and 18%
23% O2




Mann et al. (2009)

15 L Algal pond with absorption unit filled with real biogas 44%
48% CO2,55%
71% CH4 and 1% H2S




74%
95%

88%
97% CH4




Conde et al. (1993)

Column photobioreactor incorporated with real biogas has compositions—70%
72% CH4, 17%
19% CO2

Photosynthetic photon flux density—35.6 µmol photons/m2/s

100%

10%
24% O2

0.041 g/L/d

Converti et al. (2009)

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

Gas flow rate—250
2000 mL/min

85%

5.2%
6% O2




Mandeno et al. (2005)

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401

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 designing cost-effective, energy-efficient 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-of-concept 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 environmentalfriendly 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.

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401. Zhao, Y., Wang, J., Zhang, H., Yan, C., Zhang, Y., 2013. Effects of various LED light wavelengths and intensities on microalgae-based simultaneous biogas upgrading and digestate nutrient reduction process. Bioresour. Technol. 136, 461
468. Zhao, Y., Sun, S., Hu, C., Zhang, H., Xu, J., Ping, L., 2015. Performance of three microalgal strains in biogas slurry purification and biogas upgrade in response to various mixed lightemitting diode light wavelengths. Bioresour. Technol. 187, 338
345. Zhao, Y., Guo, G., Sun, S., Hu, C., Liu, J., 2019. Co-pelletization of microalgae and fungi for efficient nutrient purification and biogas upgrading. Bioresour. Technol. 289, 121656.

Chapter 15

Biogas upgrading and life cycle assessment of different biogas upgrading technologies Moonmoon Hiloidhari1 and Shilpi Kumari2 1

IDP in Climate Studies, Indian Institute of Technology Bombay, Mumbai, India, 2Centre for Energy Studies, Indian Institute of Technology Delhi, New Delhi, India

Chapter outline 15.1 Introduction 15.2 Biomethanation 15.2.1 Cleaning of biogas 15.2.2 Upgrading of biogas into biomethane 15.3 Brief overview of life cycle assessment

413 415 416 417

15.4 Life cycle assessment of biogas upgrading technologies 15.5 Conclusions References Further reading

423 441 441 445

421

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; Bedoi´c 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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00015-5 © 2021 Elsevier Inc. All rights reserved.

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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 (Bedoi´c, et al., 2019; Masebinu et al., 2014; Bond and Templeton, 2011). 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 longchain 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; De˛bowski et al., 2013; Rajendran et al., 2020). 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 terms of global warming potential (GWP) and other environmental indicators like eutrophication, acidification, ozone layer depletion, and

FIGURE 15.1 Various stages of biogas production and upgrading.

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

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

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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 passing 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

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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

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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 278.5 C at atmospheric pressure, while CH4 condenses at 2161 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 226 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

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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. The review suggests that water scrubbing technology is preferable to the PSA method and chemical absorption. Water scrubbing, and absorption with chemical TABLE 15.1 Advantages and disadvantages of biogas upgrading technologies (Luo et al., 2012; Awe et al., 2017; Bauer et al., 2013). Techniques

Advantages

Absorption

G

G

G G

Adsorption

G G

G

G

G

G

Membrane separation

G

G

G

G

Cryogenic technologies

G G

Disadvantages

Highly efficient (97% 99%) CH4 with very less loss of (0/1% 2%) of CH4 Simultaneous removal of other impurities including H2S, O2, N2 It requires less material Can be operated even at a low concentration of impurities

G

Highly efficient (98% CH4) May be applicable for small to medium plants Adsorbent can be regenerated Other impurities may be removed No waste water produced as it is a dry process No bacteria contaminant of off-gas

G

In comparison to above, it requires less investment Operation is simple and high reliable Environmentally friendly (no waste water, no air pollution) No corrosion of plant and hence, low maintenance cost

G

High CH4 (90% 98%) Low extra energy cost to reach liquid biomethane

G

G G

G G

G

G

G

G

G

High investment, operation, and power cost Difficult in operation Clogging in plant and salt precipitation Corrosion of scrubbing column Waste water produced and microbial growth may occur

High expense, heat, and power required High investment, operation, and power cost Poisoning of the amines by O2 or other chemicals

Membrane blockage or damage if H2S is present (requires pretreatment) High investment due to high membrane cost Operation is slightly difficult as a multiple-step process is required to achieve high purity and less CH4 loss High operation and maintenance costs High energy requirement for cooling

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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 economyenvironmental 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).

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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 (Guine´e, 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 end-points of an LCA study as well as the

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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 allocationbased 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 (Guine´e, 2001; ISO, 2006). It takes into account 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.

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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 1 , 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.

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. 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

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TABLE 15.2 List of software available for life cycle assessment. Software

Developer

Source

SimaPro

PRe´ 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/lcasoftware-ebalance

EIME

Bureau Veritas CODDE

www.codde.fr/en/our-software/eimepresentation

EarthSmart

EarthShift Global

www.earthshiftglobal.com/software/ earthsmart-lca-software



This is in Dutch and CML translates to English as Institute of Environmental Sciences (CML, Centrum voor Milieukunde Leiden).

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.

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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 accounts 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

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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 terms 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

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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; De˛bowski et al., 2013; Uggetti et al., 2016). Ferreira et al. (2019) assessed life cycle energy consumption and CO2 emissions of pilot- and real-scale 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 accounted 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 energy-based 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 terms 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-to-liquid plant are necessary to improve the overall environmental performance of the biogas to synthetic biodiesel system. 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

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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.

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 terms 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 terms of economic analysis through life cycle costing, the authors noted that high-pressure 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 1 . 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 terms 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

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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. 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

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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).

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 Summary of results of selected life cycle assessment studies concerning biogas upgrading technologies. Author(s)

Biogas upgrading technology

Life cycle assessment type (LCA)

Goal

Functional unit

Life cycle impact assessment method

Results and conclusions

Starr et al. (2012)

High-pressure water scrubbing (HPWS), alkaline with regeneration (AwR) and bottom ash upgrading (BABIU)

Attributional

LCA of HPWS, AwR, and BABIU biogas upgrading technologies

Removal of 1 ton of CO2

CML 2001, cumulative energy demand (CED)

AwR upgrading technology has the highest environmental impacts in all impact categories except global warming potential (GWP) due to high energy demand for alkaline reactants production. The GWP is negative for AwR and BABIU but positive for HPWS. The BABIU technology is the most environmentally friendly when compared with HPWS, AwR, BABIU, organic physical scrubbing, pressure swing adsorption (PSA), membrane separation (MS), and cryogenic separation (CS) (Continued )

TABLE 15.3 (Continued) Author(s)

Biogas upgrading technology

Life cycle assessment type (LCA)

Goal

Functional unit

Life cycle impact assessment method

Results and conclusions

Florio et al. (2019)

MS, CS, PSA, chemical scrubbing (CSc), HPWS

Attributional with system expansion

LCA of biomethane production using different technologies

Use of 1 m3 of biogas produced from biogenic waste feedstock

ReCiPe 2016 midpoint (H)

Membrane upgrading technology is slightly more environmentally superior to the other upgrading technologies. The environmental gains could be further enhanced if the digestate (coproduct of the anaerobic digestion process) is used as organic fertilizer. Future improvement in terms of CO2 recovery from upgrading technologies, more site-specific and industry data and economic evaluation of biomethane has been suggested to aid the renewable energy decision-making process

Castellani et al. (2018)

Hydrate-based biogas upgrading

Attributional

Assess primary energy consumption and GHG emissions from biosynthetic methane production via hydrate-based biogas upgrading and CO2 methanation

Unitary physical amount of produced biosynthetic methane (kg or Nm3)

IPCC 2013 GWP 100a and CED

Carbon and energy footprints of biosynthetic methane are 1.875 kg CO2eq/ Nm3 and 40.57 MJ/ Nm3. Nearly 70% of carbon footprint impact is contributed by conventional gridenergy consumption. However, when solar PV-based electrical energy is used for biosynthetic methane production, the carbon footprint and energy footprint drops to 0.7081 kgCO2eq/ Nm3 (approx. 2 60%) and 28.55 MJ/Nm3 (approx. 2 30%)

Lorenzi et al. (2019)

Solid oxide electrolyzer cells (SOECs) and HPWS

Attributional

Comparative LCA between SOECs and HPWS

Production of 1 m3 biomethane

ReciPe 2016 midpoint (H)

For 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 (Continued )

TABLE 15.3 (Continued) Author(s)

Biogas upgrading technology

Life cycle assessment type (LCA)

Goal

Functional unit

Life cycle impact assessment method

Results and conclusions

yield is higher. The source of electricity for the upgrading process and carbon content of electricity is the key factor that determines the environmental burden of biogas upgrading Ferreira et al. (2019)

Photosynthetic algae biogas upgrading units

Attributional

Life cycle energy consumption and CO2 emissions of pilot- and real-scale photosynthetic algal biogas upgrading units

MJ of CH4 produced

ReCiPe 2016 midpoint

At industrial scale, biomass dewatering and drying account 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). High impacts on climate change and low impacts on fossil and water depletion are observed for the biogas upgrading unit NavasAnguita et al. (2019)

Fischer Tropsch (FT) synthesis and dry reforming process

Attributional

LCA of synthetic fuel produced from FT synthesis and dry reforming process

Production of 1 kg of synthetic biodiesel at the plant

CML

GWP, acidification potential, and eutrophication potential of synthetic biodiesel [dry reforming of methane (DRM) biodiesel] are higher than conventional diesel. But it performs better in ozone layer depletion and cumulative nonrenewable energy demand (CEDnr) compared to conventional diesel. Thus DRM diesel is not yet a competitive alternative from an environmental perspective. Reduction (Continued )

TABLE 15.3 (Continued) Author(s)

Biogas upgrading technology

Life cycle assessment type (LCA)

Goal

Functional unit

Life cycle impact assessment method

Results and conclusions

in methane leakage, heat demand at the biogas production phase and reducing NH3 and NO emissions in the biogas-to-liquid plant are necessary to improve the environmental performance of the biogas to synthetic biodiesel system. If 100% avoidance of methane leakage in biogas production is achieved, reduction of 4.14 kg CO2eq per 1 kg of DRM diesel would be possible. This would lead to a final carbon footprint of 1.60 kg CO2eq/kg

DRM diesel combusted, a 56% saving in greenhouse gas (GHG) emissions when compared with an equivalent amount of conventional diesel Lombardi and Francini (2020)

HPWS, amine scrubbing (AS), potassium carbonate scrubbing (PCS), PSA, and membrane permeation (MP)

Attributional

Comparative LCA and life cycle cost of five biogas to biomethane upgrading technologies

Upgrading 1 Nm3 of raw biogas

CML-IA baseline

AS proved to be the best performing upgrading technology in terms of GWP, mainly due to low methane leakage and minimal electricity consumption. Considering the upgrading part only, the AS shows 13%, 1%, 23%, and 3% less GWP than HPWS, PCS, PSA, and MP, respectively. CH4 emission is the most important contributor to the HPWS, PSA, and MP processes (12.5%, 18%, and 6.5%, respectively with reference to the (Continued )

TABLE 15.3 (Continued) Author(s)

Biogas upgrading technology

Life cycle assessment type (LCA)

Goal

Functional unit

Life cycle impact assessment method

Results and conclusions

total impact of the upgrading part). In the cases of AS, the use of solvents has 42.5% contribution of the GWP, compared to 4.5% contribution to GWP for PCS process. Overall, environmental performance in the processes is strongly influenced by electricity consumption Ardolino et al. (2018)

MS

Attributional with system expansion

Assess environmental performance of biomethane (upgraded via MS process) as transportation fuel

Treatment of 100 t/d of organic waste

IMPACT 2000 1

Production of biomethane for road transportation is better than the production of energy in terms of reduction in GHG emissions and nonrenewable energy consumption

Leonzio (2016)

Chemical adsorption

Attributional

Assess environmental performance of biogas upgrading by chemical adsorption process using three different aqueous solutions of MEA, NaOH, and KOH

Generation of 1 kWh of biomethane

CML 2001

Environmental impact of the process using KOH aqueous solution is lower, followed by chemical absorption with MEA and NaOH aqueous solution

Collet et al. (2017)

Power-to-gas (P2G)

Attributional

Technoeconomic and LCA of methane production using biogas upgrading and P2G technology

Production of 1 MJ energy through methane combustion in a boiler

IPCC 2013 GWP 100a

Electricity consumption is an important contributor to GHG emissions for the P2G technologies. Continuous P2G process generates more GHGs than direct injection. Intermittent P2G operation using renewable electricity can reduce GHG emissions substantially. The GHG emissions from continuous P2G are higher than biogas upgrading, but much lower than natural gas combustion. When sewage sludge is used to produce energy, (Continued )

TABLE 15.3 (Continued) Author(s)

Biogas upgrading technology

Life cycle assessment type (LCA)

Goal

Functional unit

Life cycle impact assessment method

Results and conclusions

main GHG emissions come from digestate management of CH4 loss and emissions of N2O. Reducing electricity consumption in the electrolysis process, and integrating renewable credits from CO2 valorization can make P2G competitive in the future Zhang et al. (2017)

P2G

Attributional with system expansion

Environmental performance of P2G using LCA

For P2H: 1 MJ of hydrogen generated. For P2M (power to methane): 1 km of distance traveled by a compressed natural gas-fueled mid-size passenger vehicle

Life Cycle Data System, 2011 Midpoints

Type of electricity and sources of CO2 greatly influence the life cycle GHG emissions of P2G. Electricity supply from fossil fuels results in P2G causing higher life cycle GHG emissions than conventional technologies

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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.

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Møller, J., Boldrin, A., Christensen, T.H., 2009. Anaerobic digestion and digestate use: accounting of greenhouse gases and global warming contribution. Waste Manage. Res. 27 (8), 813 824. Montgomery, L.F., Bochmann, G., 2014. Pre-Treatment of Feedstock for Enhanced Biogas Production. IEA Bioenergy, Ireland, pp. 1 20. 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 FischerTropsch synthesis. Fuel 235, 1492 1500. Nie, H., Jiang, H., Chong, D., Wu, Q., Xu, C., Zhou, H., 2013. Comparison of water scrubbing and propylene carbonate absorption for biogas upgrading process. Energy Fuels 27 (6), 3239 3245. Okeke, I.J., Mani, S., 2017. Techno-economic assessment of biogas to liquid fuels conversion technology via Fischer-Tropsch synthesis. Biofuels Bioprod. Bioref. 11 (3), 472 487. Persson, M., Jo¨nsson, O., Wellinger, A., 2006. Biogas Upgrading to Vehicle Fuel Standards and Grid Injection, Vol. 37. IEA Bioenergy, pp. 1 34, December. Petersson, A., Wellinger, A., 2009. Biogas upgrading technologies developments and innovations. IEA Bioenergy 20, 1 19. Petrillo, A., De Felice, F., Jannelli, E., Autorino, C., Minutillo, M., Lavadera, A.L., 2016. Life cycle assessment (LCA) and life cycle cost (LCC) analysis model for a stand-alone hybrid renewable energy system. Renew. Energy 95, 337 355. Pizzol, M., Laurent, A., Sala, S., Weidema, B., Verones, F., Koffler, C., 2017. Normalisation and weighting in life cycle assessment: quo vadis? Int. J. Life Cycle Assess. 22 (6), 853 866. Rajendran, K., Mahapatra, D., Venkatraman, A.V., Muthuswamy, S., Pugazhendhi, A., 2020. Advancing anaerobic digestion through two-stage processes: current developments and future trends. Renew. Sustain. Energy Rev. 123, 109746. Reap, J., Roman, F., Duncan, S., Bras, B., 2008. A survey of unresolved problems in life cycle assessment-Part 1: goal and scope and inventory analysis. Int. J. Life Cycle Assess. 13 (4), 290 300. Ryckebosch, E., Drouillon, M., Vervaeren, H., 2011. Techniques for transformation of biogas to biomethane. Biomass Bioenergy 35 (5), 1633 1645. Sahota, S., Shah, G., Ghosh, P., Kapoor, R., Sengupta, S., Singh, P., et al., 2018. Review of trends in biogas upgradation technologies and future perspectives. Bioresour. Technol. Rep. 1, 79 88. Scholwin, F., Held, J., 2013. Biomethane from anaerobic processes. In: Kaltschmitt, M., Themelis, N.J., Bronicki, L.Y., So¨der, L., Vega, L.A. (Eds.), Renewable Energy Systems. Springer, New York. Scholz, M., Melin, T., Wessling, M., 2013. Transforming biogas into biomethane using membrane technology. Renew. Sustain. Energy Rev. 17, 199 212. Seabra, J.E., Macedo, I.C., Chum, H.L., Faroni, C.E., Sarto, C.A., 2011. Life cycle assessment of Brazilian sugarcane products: GHG emissions and energy use. Biofuels Bioprod. Bioref. 5 (5), 519 532. SoCalGas. Power-to-Gas Technology. 2020. https://www.socalgas.com/smart-energy/renewablegas/power-to-gas (accessed 20.07.20). Sridhar, S., Smitha, B., Aminabhavi, T.M., 2007. Separation of carbon dioxide from natural gas mixtures through polymeric membranes—a review. Separation. Purif. Rev. 36 (2), 113 174. Starr, K., Gabarrell, X., Villalba, G., Talens, L., Lombardi, L., 2012. Life cycle assessment of biogas upgrading technologies. Waste Manage. 32 (5), 991 999.

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Tambone, F., Scaglia, B., D’Imporzano, G., Schievano, A., Orzi, V., Salati, S., et al., 2010. Assessing amendment and fertilizing properties of digestates from anaerobic digestion through a comparative study with digested sludge and compost. Chemosphere 81 (5), 577 583. Uggetti, E., Passos, F., Sole´, M., Garc´ıa, J., Ferrer, I., 2016. Biogas from algae via anaerobic digestion. Algae Biotechnology. Springer, Cham, pp. 195 216. Walsh, J.J., Jones, D.L., Edwards-Jones, G., Williams, A.P., 2012. Replacing inorganic fertilizer with anaerobic digestate may maintain agricultural productivity at less environmental cost. J. Plant Nutr. Soil Sci. 175 (6), 840 845. Wang, Q.L., Li, W., Gao, X., Li, S.J., 2016. Life cycle assessment on biogas production from straw and its sensitivity analysis. Bioresour. Technol. 201, 208 214. Wang, C., Malik, A., Wang, Y., Chang, Y., Lenzen, M., Zhou, D., et al., 2020. The social, economic, and environmental implications of biomass ethanol production in China: a multiregional input-output-based hybrid LCA model. J. Clean. Prod. 249, 119326. Wellinger, A., Murphy, 2013. J.D. In: Baxter, D. (Ed.), The biogas handbook: science, production and applications. Elsevier. Wiley, P.E., Campbell, J.E., McKuin, B., 2011. Production of biodiesel and biogas from algae: a review of process train options. Water Environ. Res. 83 (4), 326 338. Wu, M., Zhang, W., Ji, Y., Yi, X., Ma, J., Wu, H., et al., 2017. Coupled CO2 fixation from ethylene oxide off-gas with bio-based succinic acid production by engineered recombinant Escherichia coli. Biochem. Eng. J. 117, 1 6. Wulf, C., Linssen, J., Zapp, P., 2018. Power-to-gas—concepts, demonstration, and prospects. Hydrogen Supply Chains. Academic Press, pp. 309 345. Zhang, X., Bauer, C., Mutel, C.L., Volkart, K., 2017. Life cycle assessment of power-to-gas: approaches, system variations and their environmental implications. Appl. Energy 190, 326 338. Zhang, B., Su, S., Zhu, Y., Li, X., 2020. An LCA-based environmental impact assessment model for regulatory planning. Environ. Impact Assess. Rev. 83, 106406.

Further reading Feiz, R., Ammenberg, J., 2017. Assessment of feedstocks for biogas production, part I—a multicriteria approach. Resources Conserv. Recycl. 122, 373 387. Herrmann, C., Idler, C., Heiermann, M., 2016. Biogas crops grown in energy crop rotations: linking chemical composition and methane production characteristics. Bioresour. Technol. 206, 23 35. Rafique, R., Poulsen, T.G., Nizami, A.S., Murphy, J.D., Kiely, G., 2010. Effect of thermal, chemical and thermo-chemical pre-treatments to enhance methane production. Energy 35 (12), 4556 4561. Scarlat, N., Dallemand, J.F., Fahl, F., 2018. Biogas: developments and perspectives in Europe. Renew. Energy 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 1

Department of Civil Engineering, Indian Institute of Technology Guwahati, Guwahati, India, Institute of Soil, Water and Environmental Science, Agricultural Research Organization, Israel, 3 System Biology for Biofuel Group, International Centre for Genetic Engineering and Biotechnology, New Delhi, India, 4Department of Biosciences and Bioengineering, Indian Institute of Technology Guwahati, Guwahati, India, 5Department of Environmental Science, SRM University-AP, Mangalagiri, India 2

Chapter outline 16.1 Introduction 448 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 support to AD industry growth 456 16.4 Decision-support system for biomethane implantation with techno-economic analysis and policies 461 16.5 Conclusion 462 References 463

Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00016-7 © 2021 Elsevier Inc. All rights reserved.

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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. Biogasbiomethane 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

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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 (h0.76/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.

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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 (Buˇsi´c 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).

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Carbon dioxide separation from methane can also be carried out using different membrane types such as polymeric, ceramic, inorganic, cross-linked, metallic-based, 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 h99/ 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.

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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 support 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 h20/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 agrowastes. The program focused on research and development, construction and monitoring biogas plants, and spreading awareness about anaerobic digestion. Economic support 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.

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.

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TABLE 16.1 Subsidy rate for different states and categories in India (MNRE, 2019). S. no.

Central financial assistance (CFA) and states/regions and categories

Small biogas plants (Rs/m3 digester volume)

A.

Central subsidy rates applicable

1 m3

2 6 m3

1.

Northeastern (NE) region states, Sikkim (except plain areas of Assam) and including scheduled caste (SC) and scheduled tribes (ST) Categories of NE Region States

15,000

17,000

2.

Plain areas of Assam

10,000

11,000

3.

Jammu & Kashmir, Himachal Pradesh, Uttarakhand, Nilgiri of Tamil Nadu, Sadar Kurseong and Kalimpong Sub-Divisions of Darjeeling, Sunderbans and Andaman and Nicobar Islands

7000

11,000

4.

Scheduled castes/Scheduled Tribes of all other States except NE Region States (including Sikkim)

7000

11,000

5.

All others

5500

9000

B.

Turn-key job fee including warranty for 5 years and quality control (in Rs. per plant)

Rs.1500/- per plant for fixed dome Deenbandhu type and floating gasholder KVIC type brick masonry models

C.

Additional subsidy (CFA) for toiletlinked biogas plants (in Rs. per plant)

1200/-

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

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(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). Germany, 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.

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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 support 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).

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16.3.6 Policy instruments introduced in various countries as a support 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 Germany Germany 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 support the biogas producers and consumers. Through the Biofuels Quota Act of 2007, Germany 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. Germany 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, Germany 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. Germany 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 Germany 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).

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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 support 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 support 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 support 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

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Growth Strategy was announced by the UK government under which d2.5 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 support 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 h375 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 premium 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

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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 supported 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 support 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 support 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

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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 support biogas production are listed in Table 16.2. TABLE 16.2 Key policies applied in Denmark, Austria, and the Netherlands to support biogas production. Country

Policy

Detail

Denmark

1993 FIT policy

Constant rate of 4.3 hct/kWh for AD industry

1999 Renewable portfolio standard with a system of tradable green certificates

Before the expiry of the 1993 FIT policy, new legislature, tradable green certificates provided in place of fixed rate, biogas industry growth stagnation

2004 FIT policy and its amendment in 2009

Increased rates of 8 hct/kWh for a period of 20 years (EREC, 2009), biogas production increased from 55 to 67 ktoe within 4 years (2004 2008); amendment in 2009; rates made variable, for instance average payment including market price and bonus would reach 9.7 hct/kWh under the new system (Ferreira et al., 2012)

The Green Growth Agreement, 2009

Target to use 50% of livestock manure for green energy by 2020 (Bundgaard et al., 2014)

Energy Strategy, 2050 (2011)

100% renewable energy sources by 2050 (across all energy sectors); reduction of GHE by 40% by 2020 with respect to 1990 levels (Ropenus and Jacobsen, 2015)

Estrada et al. (2012)

Target to reduce CO2 emissions by 34% with respect to 1990; 35% share from renewable sources in the total energy consumption (Ropenus and Jacobsen, 2015) (Continued )

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TABLE 16.2 (Continued) Country

Policy

Detail

Austria

1998 FIT policy (amended 2002, 2006, 2010)

Constant rate of 5 hct/kWh for 13 years (Bangalore et al., 2016) 2002 amendment, 13.4 hct/kWh (Gabauer et al.2008), biogas production increased from 8 to 140 ktoe within 4 years from 2002 to 2006 (Eurostat, 2012) 2006 amendment, 14.2 hct/kWh for 10 years (Braun et al., 2008), unable to compete with the increased energy crops prices. Biogas production became stagnant at 140 ktoe (Eurostat, 2012) 2010 amendment, 15.8 hct/kWh for 15 years (Drosg, 2011)

National Green Electricity act, 2002 (amended 2006, 2007, 2008)

Financial support to electricity generated via renewable sources, additional bonuses for combined heat and electricity generation from renewable technology (Nachmany et al., 2015)

2003 MEP (Environmental Quality of Electricity Production)

7.7 hct/kWh for 10 years, biogas production increased from 27 to 47 ktoe within 3 years from 2003 to 2006 (Bangalore et al., 2016)

2006 SDE (new stimulation program)

11.2 hct/kWh offered for a period of 12 years, biogas production increased from 47 to 207 ktoe within 4 years from 2006 to 2010 (Eurostat, 2012)

2011 SDE 1

Feed-in premium (FIP) subsidy for 15 years, covers the difference between wholesale market prices of electricity and the cost price of electricity from renewable sources (IEA, 2011)

Netherlands

16.4 Decision-support 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

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evaluation at the supply chain level of fossil and renewable energy sources is lacking. To that end, an Excel-based mathematical model, called the Decisionsupport 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-support 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 techno-economic analysis of biomethane generation was developed by Billig and Thra¨n (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-support 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-support system is developed based on multicriterion information, including technical, social, and environmental.

16.5 Conclusion To commercialize biomethane technologies, policy support is necessary. Some of the successful policies implemented in Europe could be used in other countries as well for efficient commercialization. Dynamic support 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 support will encourage the industry toward innovation, thus reducing the cost. An integrative decision-support 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,

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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.

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Chapter 17

Large-scale biogas upgrading plants: future prospective and technical challenges Ram Chandra Poudel1, Dilip Khatiwada2, Prakash Aryal3 and Manju Sapkota4 1

Department of Biological Sciences, University of Bergen, Bergen, Norway, 2Division of Energy Systems, Department of Energy Technology, KTH Royal Institute of Technology, Stockholm, Sweden, 3Department of Chemical Engineering, Monash University, Clayton, VIC, Australia, 4 Institute of Chemistry, Bioscience and Environmental Engineering, Faculty of Science and Technology, University of Stavanger, Stavanger, Norway

Chapter outline 17.1 Introduction 467 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 Emerging Technologies and Biological Systems for Biogas Upgrading. DOI: https://doi.org/10.1016/B978-0-12-822808-1.00017-9 © 2021 Elsevier Inc. All rights reserved.

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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 laboratory-scale 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 hydrogen-mediated 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.

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

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FIGURE 17.1 Different types of feedstock for large-scale biogas plants (IEA Bioenergy, 2017).

TABLE 17.1 Constituents of biogas from different waste-to-biogas conversion options, and their consequences (Awe et al., 2017; Sun et al., 2015; Bailo´n and Hinge, 2012; Ryckebosch et al., 2011). Biogas constituents

Biogas (biomass)

Biogas (waste water/ sludge)

Biogas (landfill waste)

Possible consequences

CH4 (vol.%)

60
70

55
75

35
65

Low methane concentration has lower Wobbe Index

CO2 (vol.%)

30
40

19
45

15
40

Decreases the energy content (calorific value or Wobbe Index) of gas and increases compression and transportation cost

N2 (vol.%)

0.2

0
2

15

Decreases the energy content (calorific value or Wobbe Index) of gas (Continued )

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PART | IV Policy implications for biogas upgrading

TABLE 17.1 (Continued) Biogas constituents

Biogas (biomass)

Biogas (waste water/ sludge)

Biogas (landfill waste)

Possible consequences

O2 (vol.%)

0

0
1

1

Favors aerobic microorganisms to grow and cause explosion at a certain ratio of CH4 and O2

H2 (vol.%)

0

Nf

0
3

Decreases the energy content (calorific value or Wobbe Index) of gas

Hydrocarbons total (vol.%)

0

Nf

0

Upon combustion forms toxic compounds (increases the calorific value)

NH3 (ppm)

100

Nf

5

Corrosive when reacts with water

H2S (ppm)

0
4000

63
3000

0
100

Health hazard (poisonous); corrosive reaction with water and forms sulfuric acid when combusted and damages pipelines, CHP units, gas storage tanks, etc.

Siloxanes (ppmv)

0
0.4

1.5
15

0.1
4

Combustion produces silicon dioxide (silica) which gets deposited onto the CPH unit, pipelines, metal instruments, etc. and causes overheating, abrasion, and damage to engines and instruments

H2O vapor (mg/N m3)

1
5a

a

1
5a

Lowers the energy contents and reacts with H2S, NH3, and CO2 to produce corrosive acids

Total chlorine (mg/N m3)

100

Nf

5

Corrosion of combustion engines

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. a Saturation at biodigester temperature.

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(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 (Bailo´n 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 Comparison of biomethane standards (gas grid injection and vehicle fuel) in different countries. Parameters

Sweden (vehicle fuel) (Awe et al., 2017; Bailo´n and Hinge, 2012)

Germany (gas grid) (Awe et al., 2017; Rodero et al., 2018)

The Netherlands (gas grid) (Awe et al., 2017; Rodero et al., 2018)

North Sea natural (gas grid) (Awe et al., 2017; Rodero et al., 2018)

Lower heating value (MJ/N m3)

Nf

30.2
47.2

31.6
38.7

40

Higher (H)

44.7
46.4

46.1
56.5

43.46
44.41

55

Lower (L)

43.9
47.3

37.8
46.8

Nf

Nf

Methane (vol.%)

$97

98.3

$85

85
92

CO2 (vol.%)

#3

#6 (dry)

#6 (10
10.3; regional grid)

0.2
1.5

O2 (vol.%)

#1

#3

#0.5

0

N2 (vol.%)

Nf

Nf

Nf

0.3
1.0

Heavy hydrocarbons (vol.%)

Nf

Nf

Nf

9

H2 (vol.%)

#0.5

#5

,12

0

H2S (ppmv)

#15.2

#5

#5

1.1
5.9

Total sulfur (ppmv)

,23

#30

,45

-

WI

(Continued )

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PART | IV Policy implications for biogas upgrading

TABLE 17.2 (Continued) Parameters

Sweden (vehicle fuel) (Awe et al., 2017; Bailo´n and Hinge, 2012)

Germany (gas grid) (Awe et al., 2017; Rodero et al., 2018)

The Netherlands (gas grid) (Awe et al., 2017; Rodero et al., 2018)

North Sea natural (gas grid) (Awe et al., 2017; Rodero et al., 2018)

NH3 (ppmv)

#20

#20

#3

0

Mercaptans (ppmv)

Nf

#15

#10

—

Halogenated compounds (mg/N m3)

#1

#1

#75

0

Siloxanes (ppm)

Nf

Nf

#5

Nf

H2O dew point ( C)

#25 at AT #29 at 200 bar

ground temperature

#10 at 8 bar

Nf

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.

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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; Mun˜oz 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). 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 passes through it while CH4 is retained by the membrane. This process

FIGURE 17.2 Number of physicochemical upgrading technologies in 2017 (EBA, 2018).

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PART | IV Policy implications for biogas upgrading

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 278 C and 160 C, respectively (Awe et al., 2017; Thra¨n UFZ et al., 2014). When the gas mixture is cooled at 285 C to 2110 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.

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 Germany 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; Ko¨tter 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

TABLE 17.3 Technical parameters of different physicochemical biogas upgrading processes. Technology

CH4 recovery (%)

CH4 loss (%)

H2S (ppmv)

CO2 (%)

Pressure (bar)

Temperature ( C)

References

Water scrubbing

.97

,2

,2

,2

6
12

40

Adnan et al. (2019), Awe et al. (2017), Niesner et al. (2013)

Organic physical (solvent) scrubbing

96
98.5

,2

7
8







55
80

Adnan et al. (2019), Awe et al. (2017)

Chemical (amine) scrubbing

.99

,0.05

,4

,1

1
2

20
65

Adnan et al. (2019), Prussi et al. (2019), Niesner et al. (2013)

Pressure swing adsorption

98

2
4




1
2

4
10




Niesner et al. (2013), Ryckebosch et al. (2011)

Membrane separation

99.5 (MSS) 90
97 (SSS)

,0.5




1
3

20
36 (MSS) 1 atm. (SSS)




Niesner et al. (2013), Scholz et al. (2013)

Cryogenic separation

90
98

,2







200

B225, 2110

Adnan et al. (2019), Ryckebosch et al. (2011)

MSS, Multistage system; SSS, single-stage system.

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PART | IV Policy implications for biogas upgrading

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; MendozaHernandez 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 two-phase 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. 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 Germany. 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 joint 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 egas) above 96% (Bailera et al., 2017; Benjaminsson et al., 2013).

TABLE 17.4 Different catalytic methanation upgrading technologies with technical parameters. Catalytic methanation

Scale

ETOGAS and Audi e-gas

Commercial

MeGa-StoRE

Reactor type

Electrolyzer

CH4 content (%)

Output gas (N m3/h)

Tempr ( C)

Pressure (bar)

References

Type

EI (kWe)

Isothermal fixed bed

Alkaline

6000

.96

325

250
550

8
10

Bailera et al. (2017), Benjaminsson et al. (2013), Sveinbjo¨rnsson et al. (2017)

Pilot

Air cooled




Bottled H2

.97

1

270

8

Bailera et al. (2017), ForskEl (2015), Høegh Jensen (2015)

Haldor Topsøe

Pilot

Reactor with steam

SOEC

50

$ 96

10

280

Medium

Benjaminsson et al. (2013), ForskNG (2011), Sveinbjo¨rnsson et al. (2017)

RENOVA GAS

Demonstration

Multichannel

Alkaline

15

2

275
330

25

Bailera et al. (2017)

HELMETH

Pilot

Isothermal

SOEC

30

300

15

Bailera et al. (2017)

94.1


r

EI, Electricity input; MeGa-StoRE, methane gas storage of renewable energy; SOEC, solid oxide electrolyzer cell; Temp , temperature.

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PART | IV Policy implications for biogas upgrading

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; Mar´ın et al., 2018). This technology offers the unique opportunity to utilize CO2 from biogas as a feedstock either for biomethane

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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 (Mun˜oz 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; Stru¨bing 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; D´ıaz 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). 4H2 1 CO2 -CH4 1 2H2 O

GO 5 2 165 kJ=mol

ð17:1Þ

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

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PART | IV Policy implications for biogas upgrading

rise to CH4 with optimal production (Luo et al., 2014; Stru¨bing 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. 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

TABLE 17.5 Chemoautotrophic methanation technologies and their technical specifications. Chemoautotrophic methanation

Scale

Reactor type

Electrolyzer

Electrochaea (BioCat methanation) MicrobEnergy (BioPower2Gas)

Technology

Energy input (kWe)

Commercial

Ex situ

Alkaline

Pilot

In situ 1 ex situ

PEM

PEM, Polymer electrolyte membrane.

CH4 content (%)

Operating Temp ( C)

Pressure (bar)

1000

.99

63

10

Bailera et al. (2017), Benjaminsson et al. (2013), Electrochaea (2018)

300

.95

40

5

Bailera et al. (2017), Benjaminsson et al. (2013)

References

482

PART | IV Policy implications for biogas upgrading

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, Germany, 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

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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 membrane-bound 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): Direct pathway: CO2 1 8H1 1 8e2 -CH4 1 2H2 O

ð17:2Þ

Indirect pathway: 8H1 1 8e2 -4H2

ð17:3Þ

CO2 1 4H2 -CH4 1 2H2 O

ð17:4Þ

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, redesigning the bioreactor by removing the

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membrane separation of the anode from the cathode will make the reactor cost-effective 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 (Mar´ın 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. (Alc´antara et al., 2015; Rodero et al., 2019; ToledoCervantes 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 m3d21 per m3 reactor volume. Likewise, a first demonstration-scale (proof-of-concept) of photosynthetic biogas upgrading in an algal-bacterial

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photobioreactor showed that CH4 enrichment efficiency was B90%, 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. Hydrogen-mediated 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 massliquid 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

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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.

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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 contactors, 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 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 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

493

494

Index

Alkaline water electrolysis (AWE), 189 192 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, 413 417, 467 468 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 reactor configurations, 210 211 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

Index 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 absorptionmembrane, 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 cryogenicmembrane 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 reactor configurations, 330 337 standards for natural gas quality, 322t Biomethanation, 415 421, 448 449. See also Catalytic/thermochemical methanation; Chemoautotrophic methanation cleaning of biogas, 416 417

495

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-support 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

496

Index

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, 32 33, 93 94, 223 225, 321 322, 346 347, 363 364, 413 414, 467 468 biomethanation, 479 CO2R process, 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), 104 106, 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 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, 386 387, 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

Index 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 (CPB 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

497

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 support methodology (DSM), 461 462 Decision-support for Digester-Algae IntegRation for Improved Environmental and Economic Sustainability (DAIRIEES), 461 462 Decision-support 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), 206 208, 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

498

Index

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, 342 343, 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 biomethanation, 332 337 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 ˚ sieve, 106 5A 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

Index Flemion, 194 Flooding point of packing, 41 42 Flow rate, 103 Fluidized bed methanation reactors, 200 202 design, 200 developments, 201 202 reactor configurations, 200 201 Foaming, 44 45 Free volume of polymer membrane, 130 Freezing methods, 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 contactors, 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

499

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 contactors, 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, 223 225, 236, 295 296, 321 322, 363 364, 383 384, 413 416, 467 468 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

500

Index

Hydrogen sulfide (H2S) (Continued) 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 11, 414 Hydrolytic bacteria, 328 329

I Ideal membrane selectivity, 121 Immobilized liquid membranes, 306 Impurities removal, 416 417 In situ biomethanation, 331 332 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

Index contactors, 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

501

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 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, 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 diffuser, 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

502

Index

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, 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, 413 416, 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, 413 416, 467 468 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 support 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, 121 122 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

Index 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, 304 306 Physicochemical upgrading technologies, 472 474, 475t Picochlorum sp., 394 Piperazine (PZ), 31 33, 65 66 Plasticization, 140 Plugging, 86 87 Poiseuille diffusion, 110 112 Poiseuille diffusivity coefficient, 110 111 Policy support, 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 adsorption into, 274 276 Polymer electrolyte fuel cells (PEFCs), 189 190 Polymer electrolyte membrane (PEM), 474 476 electrolysis, 324 electrolyzers, 193 196, 193f design, 193 196 developments, 195 196 reactor configurations, 194 195 water electrolysis, 193 194

503

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), 190 192, 428 429 Potassium iodide (KI), 301, 303 304 Potassium titanate (K2TiO3), 192 Power consumption, 100 Power to gas (PtG), 187 188, 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

504

Index

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 support (RP support), 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 support 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 adsorption into, 266 268 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), 190 192, 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 developments, 197 198 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

Index Solution-diffusion mechanism, 122 Solvent selection, 32 33 Spirulina, 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 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 CP9, 314 T. thioparus, 309 Thiomicrospira denitrificans, 309 THIOPAQ O&G process, 310 THIOPAQ process, 310 311 Three-phase reactor, 202 203

505

design, 202 reactor configurations, 202 203 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), 255 257, 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

506

Index

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, 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 adsorption into, 268 271 classification and parameters, 270t physical parameters, 271t Zero-gap cell, 193 Zinc oxide (ZnO), 301 ZSM-5, 106