Bioelectrosynthesis: Principles and Technologies for Value-Added Products 3527343784, 9783527343782

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Bioelectrosynthesis: Principles and Technologies for Value-Added Products
 3527343784, 9783527343782

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Bioelectrosynthesis

Bioelectrosynthesis Principles and Technologies for Value-Added Products

Edited by Aijie Wang Wenzong Liu Bo Zhang Weiwei Cai

Editors Prof. Aijie Wang

Res. Ctr. for Eco-Environmental Sciences Key Laboratory of Environmental Biotechnology 18 Shuangqing Road Haidian District 100085 Beijing China

All books published by Wiley-VCH are carefully produced. Nevertheless, authors, editors, and publisher do not warrant the information contained in these books, including this book, to be free of errors. Readers are advised to keep in mind that statements, data, illustrations, procedural details or other items may inadvertently be inaccurate. Library of Congress Card No.:

Dr. Wenzong Liu

Res. Ctr. for Eco-Environmental Sciences Key Laboratory of Environmental Biotechnology 18 Shuangqing Road Haidian District 100085 Beijing China Dr. Bo Zhang

Res. Ctr. for Eco-Environmental Sciences Key Laboratory of Environmental Biotechnology 18 Shuangqing Road Haidian District 100085 Beijing China

applied for British Library Cataloguing-in-Publication Data

A catalogue record for this book is available from the British Library. Bibliographic information published by the Deutsche Nationalbibliothek

The Deutsche Nationalbibliothek lists this publication in the Deutsche Nationalbibliografie; detailed bibliographic data are available on the Internet at . © 2020 Wiley-VCH Verlag GmbH & Co. KGaA, Boschstr. 12, 69469 Weinheim, Germany

Dr. Weiwei Cai

Harbin Institute of Technology State Key Lab. of Urban Water Resources No 83 Huanghe Road 150090 Harbin China

All rights reserved (including those of translation into other languages). No part of this book may be reproduced in any form – by photoprinting, microfilm, or any other means – nor transmitted or translated into a machine language without written permission from the publishers. Registered names, trademarks, etc. used in this book, even when not specifically marked as such, are not to be considered unprotected by law. Print ISBN: 978-3-527-34378-2 ePDF ISBN: 978-3-527-34384-3 ePub ISBN: 978-3-527-34381-2 oBook ISBN: 978-3-527-34382-9 Cover Design Tata Consulting Services Typesetting SPi Global, Chennai, India Printing and Binding

Printed on acid-free paper 10 9 8 7 6 5 4 3 2 1

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Contents Preface xiii

Section I Principle and Products Overview of Bioelectrosynthesis 1 1

Principle and Product Overview of Bioelectrosynthesis 3 Fang Zhang, Yuquan Wei, and Guanghe Li

1.1 1.2 1.3 1.4 1.4.1 1.4.2 1.4.3 1.4.4 1.4.5 1.5 1.5.1 1.5.2 1.6

Introduction 3 Evolution of Bioelectrosynthesis 6 Fundamental Principles of Bioelectrosynthesis 9 Plethora of Applications for Chemical Production 11 Hydrogen Production 11 Methane Production 12 Alcohol Production 16 Short-chain Organic Acid Production 17 Ammonia Production and Nitrogen Recovery 23 Key Factors for Improving MES Performance 26 Electron Transfer from the Cathode to the Cell 26 Cathode Materials 27 Summary 29 References 29

Section II Biogas Production and Upgrading Technology via Bioelectrolysis 39 2

Hydrogen Production from Waste Stream with Microbial Electrolysis Cells 41 Defeng Xing, Yang Yang, Zhen Li, Han Cui, Dongmei Ma, Xiaoyu Cai, and Jiayu Gu

2.1 2.1.1 2.1.2

Construction of MEC and Scale-up 42 Laboratory-Scale MEC 44 Pilot-Scale MEC 46

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2.2 2.2.1 2.2.2 2.2.2.1 2.2.2.2 2.2.2.3 2.3 2.3.1 2.3.2 2.3.3 2.3.4 2.4 2.4.1 2.4.2 2.4.3 2.4.3.1 2.4.3.2 2.4.3.3 2.5 2.5.1 2.5.2 2.5.3 2.5.4 2.6

Electrode Material of MEC 47 Anode of MEC 47 Cathode of MEC 49 Cathode Base Materials in MEC 49 Cathode Catalysts in MEC 49 Biological Catalysts in MEC 51 Effect of Operation Conditions on Hydrogen Production 51 Effect of Substrate on Hydrogen Production 51 Effects of Applied Voltage and Magnetic Field on Hydrogen Production 52 Effect of pH on Hydrogen Production 54 Effect of Temperature on Hydrogen Production 54 Electroactive Biofilm Microbiome and Syntrophic Interaction in MEC 54 Anodic EAM and Biofilm Formation 55 EAM in the Cathode 56 Microbial Community and Syntrophic Interaction 58 Pure Culture and Mixed Culture 58 Microbiome in Electroactive Biofilms 58 Suppressing the Methanogens 59 Coupled System for Biohydrogen Production 60 MEC–MFC-Coupled System for Biohydrogen Production 60 AD–MEC-Coupled System for Hydrogen Production 60 Solar-Powered MEC-Coupled System for Hydrogen Production 61 Other Modified MEC System for Hydrogen Production 61 Challenges and Outlook 62 Acknowledgment 63 References 64

3

A Promising Strategy for Renewable Energy Recovery: Conversion of Organic Wastes to Methane via Electromethanogenesis 71 Zhiqiang Zhao and Yaobin Zhang

3.1 3.2 3.3 3.3.1 3.3.2 3.4 3.4.1 3.4.2 3.4.3 3.5

Introduction 71 Advances in Electromethanogenesis 72 Mechanisms of Electromethanogenesis 75 Electron Transfer from Electrode to Methanogens 75 Microbial Communities of Biocathode 77 Applications of Electromethanogenesis 81 Renewable Energy Storage 81 Biogas Upgrading 82 Organic Waste Treatment 83 Outlook 86 References 87

Contents

4

Microbial Electrolysis Cell (MEC): An Innovative Waste to Bioenergy and Value-Added By-product Technology 95 Abudukeremu Kadier, Najeeb K. N. Al-Shorgani, Dipak A. Jadhav, Jayesh M. Sonawane, Abhilasha S. Mathuriya, Mohd S. Kalil, Hassimi A. Hasan, and Khulood Fahad Saud Alabbosh

4.1 4.2

Introduction 95 Microbial Electrolysis Cell (MEC) for Hydrogen Production and Waste Treatment 96 Working Principles 96 Advantages of MEC Over Other Potential Waste Treatment Technologies 97 Different Types of Waste Feedstocks Used in MECs 99 Simple or Defined Substrates 99 Glucose 106 Acetate 106 Glycerol 106 Proteins 107 Volatile Fatty Acids (VFAs) 107 DF Effluents and Other Pure Substrates 108 Wastewater Feedstocks 108 Domestic Wastewater (DW) 108 Industrial Wastewater 109 Complex or Lignocellulosic Biomass Materials 110 Waste-Activated Sludge (WAS) 110 Agricultural Wastes and Landfill Leachate 113 Current Applications of MEC 113 Hydrogen Production and Ammonium Recovery from Urine 113 Metal Removal or Recovery from Wastes 115 Existing Challenges and Bottlenecks for the Use of Wastewaters as Substrates in MECs 117 Conclusion and Future Outlook 118 Acknowledgments 119 References 119

4.2.1 4.2.2 4.3 4.3.1 4.3.1.1 4.3.1.2 4.3.1.3 4.3.1.4 4.3.1.5 4.3.1.6 4.3.2 4.3.2.1 4.3.2.2 4.3.3 4.3.4 4.3.5 4.4 4.4.1 4.4.2 4.5 4.6

5

Methane Production at Biocathodes: Principles and Applications 129 Dandan Liu, Marco Zeppilli, Marianna Villano, Cees Buisman, and Annemiek ter Heijne

5.1 5.2 5.2.1 5.2.2 5.2.3 5.2.4

Introduction 129 Fundamentals of Methane-Producing Biocathode 131 Cathode Potential and Mechanism of Methane Production 133 Methane Production Rate 135 Current-to-Methane, Voltage, and Energy Efficiencies 136 Electron Donor for Methane-Producing BESs 138

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5.3 5.3.1 5.3.2 5.3.3 5.4 5.4.1 5.4.2 5.5 5.6

Enhancing Methane Production Rates in AD 139 AD + BES Combination 139 BES + AD Combination 143 BES Within AD Combination 143 Upgrading of Biogas 145 Fundamental Aspects of Biogas Upgrading 145 Alkalinity Generation in BES Biocathodes and CO2 Removal 147 Storage of Renewable Energy Through Methane-Producing Bioelectrochemical System 150 Conclusions and Outlook 153 References 154

Section III Organic Production in Microbial Electrosynthesis System 161 6

Organic Synthesis on Cathodes 163 Annie Modestra Jampala, Sai K. Butti, and Srinivasulu Reddy Venkata Mohan

6.1 6.1.1 6.1.2 6.2 6.2.1 6.2.2 6.3 6.3.1 6.3.2 6.4 6.5

Carbon Reduction for Organics Synthesis at Cathode Gas Fermentation 164 Microbial Electrosynthesis (MES) 165 Acetate Synthesis 168 Biochemistry of Acetate Synthesis 168 Bacteria for Acetate Synthesis 171 Formic acid Synthesis 171 Direct and Indirect Conversion 172 Production Yields and Optimizations 172 Alcohol Synthesis 174 Conclusions and Future Outlook 175 Acknowledgments 176 References 176

Section IV

163

Chemical Products and Nitrogen Recovery 183

7

Inorganic Compound Synthesis in Bioelectrochemical System: Generation Rate Increase and Application 185 Lei Gao, Xi-Qi Li, Ling Wang, Wen-Zong Liu, and Ai-Jie Wang

7.1 7.2

Introduction 185 Hydrogen Peroxide Produced in BES: Optimization and Application 186 Electrode Optimization Design 188 Membrane Material Selection 189 Cation Exchange Membrane 192 Anion Exchange Membrane 193 Other Common Membranes 193

7.2.1 7.2.2 7.2.2.1 7.2.2.2 7.2.2.3

Contents

7.2.3 7.2.3.1 7.2.3.2 7.2.3.3 7.2.4 7.2.5 7.3 7.3.1 7.3.2 7.3.3 7.4 7.4.1 7.4.2 7.4.3 7.4.4 7.5 7.5.1 7.5.2 7.5.3 7.5.4 7.6

Operation Condition Optimization 194 Buffer Solution 194 Hydraulic Retention Time 195 Applied Voltage 195 Application of H2 O2 Production in BES 196 Summary 197 Metal Ion Reduction in BES: Waste Treatment and Metal Reuse 197 Metal Waste Treatment 197 Metal Reuse 198 Summary 199 Struvite Crystallization Recovery: Principle and Application in BES Systems 199 Principle of Struvite Crystallization Recovery 200 Struvite Crystal Recovery Applied in MFC 200 Struvite Crystal Recovery Applied in MEC 201 Summary 201 Ammonia Recovery and Other Inorganics Synthesis in BES Systems 202 Migration of NH4 + in BES Systems 202 Ammonia Recovery in BES Systems 202 Other Inorganics Synthesis in BES Systems 203 Summary 205 Outlook 205 Acknowledgments 206 References 206

8

Bioelectrochemical Ammonium Production – Nitrogen Removal and Recovery in BES 217 Guoqiang Zhan

8.1 8.2 8.3 8.4

Ammonium Migration and Recovery 218 Anodic Ammonium Oxidation 220 Nitrification/Denitrification in BESs 224 Existing Problems and Challenges 227 References 227

9

Bioelectrochemical Systems for Heavy Metal Pollution Control and Resource Recovery 233 Bo Zhang, Wentao Jiao, and Heming Wang

9.1 9.1.1

Introduction 233 Brief Review of Commonly Used Technologies for Heavy Metal Pollution Control and Their Respective Limitations 233 Control of Heavy Metal Pollution Through (Bio)Electrochemical Processes 234 BES and its Application in Heavy Metal Pollution Control 236 Configuration of BES 236 BES Application in Treating Heavy Metal Laden Wastewater 238

9.1.2 9.2 9.2.1 9.2.2

ix

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Contents

9.2.2.1 9.2.2.2 9.2.3 9.3

BES with Abiotic Cathode 238 BES with Biocathode 244 BES Application in Controlling Heavy Metal Polluted Soils 246 Outlook and Concluding Remarks 247 Acknowledgments 248 References 248 Section V External Electron Transfer and Electrode Material Promotion 255

10

External Electron Transfer and Electrode Material Promotion 257 Fanghua Liu, Hengduo Xu, and Jiajia Li

10.1 10.1.1 10.1.2 10.2 10.2.1 10.2.2 10.3 10.3.1 10.3.1.1 10.3.1.2 10.3.1.3 10.3.2 10.3.3 10.4 10.5

External Electron Transfer 258 Direct Electron Transfer 259 Indirect Electron Transfer 260 Promotion of Material Development 262 Carbon-Based Electrodes for Bioelectrosynthesis 263 Metal-Based Electrode for Bioelectrosynthesis 264 Modified Electrodes for High Bioelectrosynthesis 265 Electrode Modification with Carbon-Based Materials 265 Carbon Nanotube 265 Graphene 266 Activated Carbon 269 Electrode Decoration with Metal-Based Materials 269 Electrode Decoration with Other Materials 270 Interspecies Electron Transfer Pathway 272 Future Perspectives 273 References 274

11

External Electron Transfer: Pathway, Mechanism, and Microorganisms Involved 281 Cong Huang, Jun Nan, and Aijie Wang

11.1 11.1.1 11.1.2 11.2 11.3 11.3.1 11.3.2 11.3.3

External Electron Transfer of Cathode 281 Interspecies Electron Transfer (IET) 282 Direct Electron Transfer (DET) 283 Promotion of Material Development 284 Interspecies Electron Transfer Pathway 287 Mechanism of MIET 289 Mechanism of DIET 289 IET Microorganisms 290 References 290

12

Extracellular Electron Transport of Electroactive Biofilm 295 Xu Zhang

12.1 12.1.1

Electroactive Bacteria 295 Role of Multiheme Cytochromes in Extracellular Electron Transport (EET) 295 Electron Transport Across Geobacter(−Dominated) EABs 297

12.2

Contents

12.2.1 12.2.2 12.2.3

“Metallic-like” Conductivity via Microbial Nanowires 297 Redox Conduction 298 Basic Electrochemical Characterization of Redox Conductors 300 References 302 Section VI

The Microbiology of Bioelectrosynthesis 307

13

Microbial Growth and Ecological and Metabolic Characteristics in Bioelectrosynthesis Systems 309 Qian Liu and Sihao Lv

13.1 13.1.1 13.1.2 13.1.3 13.1.3.1 13.1.3.2 13.1.3.3 13.1.3.4 13.1.4 13.2

Microbial Growth Kinetics and Energetics 309 Stoichiometry of Microbial Growth Systems 309 Electrode-Respiring Bacteria Kinetics 312 MES-Associated Carbon Fixation Pathways 315 Wood–Ljungdahl Pathway 315 Reverse Tricarboxylic Acid Cycle 316 3-Hydreoxypropionate/4-Hydroxybutyrate Pathway 316 Calvin–Benson–Bassham Cycle 317 Bacterial Energetics 317 Microbial Ecological Characterization and Biofilm-Related Aspects 318 Model Electroactive Microorganisms 318 Electroactive Microorganism’s Ecology Characteristics 319 Environmental Characteristics 319 Metabolic Characteristics 320 Microbial Biofilm Formation and Characteristics 320 Electron-Release Anodic Biofilms 320 Cathodic Biofilm for Hydrogen Production 321 Cathodic Biofilm for Methane Production 321 Cathodic Biofilm for Organic Acid Production 321 Meta-omics Characterization 322 Metagenomics of MES-Associated Microorganisms 322 Metatranscriptomics of MES-Associated Microorganisms 322 Influence of Bioelectrochemistry on Microbial Community and Metabolism Pathway 323 Influence of Bioelectrochemistry on Microbial Community 323 Effects of Electrochemistry on Microbial Activity 324 Influence of Bioelectrochemistry on Microbial Metabolism Pathway 324 References 325

13.2.1 13.2.2 13.2.2.1 13.2.2.2 13.2.3 13.2.3.1 13.2.3.2 13.2.3.3 13.2.3.4 13.2.4 13.2.4.1 13.2.4.2 13.3 13.3.1 13.3.2 13.3.3

14

An Update Perspective of Electron Transfer in Electrosyntrophic Methanogenesis: From VFAs to Methane 333 Weiwei Cai, Linna Cai, and Hong Yao

14.1 14.2

Introduction 333 Interspecies Hydrogen/Formate Electron Transfer and Transport/Flow in Methanogens 333 Beyond Hydrogen/Formate Electron Carriers 336

14.3

xi

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Contents

14.3.1 14.3.2 14.4 14.5 14.5.1 14.5.1.1 14.5.1.2 14.5.1.3 14.5.1.4 14.5.1.5 14.5.1.6 14.5.2 14.6 14.7 14.8 14.8.1 14.8.2 14.9

Direct Interspecies Electron Transfer for Acetotrophic Methanogens 336 Novel Electron Donor for Hydrogenotrophic Methanogens 337 Power Drives Interspecies Electron Transfer (Kinetics and Energetics) 339 Multi-VFA Degradation Disturbance by Electrosyntrophic DIET 343 Mechanism of the Methanogenic Degradation of VFAs 344 Process of Anaerobic Digestion in an Electrochemical System 344 Conversion of Propionate and Acetate 346 Conversion of Butyrate and Acetate 347 Conversion of Formate and Acetate 348 Conversion of Valerate 349 Conversion of Mixed-VFAs vs. Individual VFA 350 DIET Process During the Methanogenesis 350 Overview of Application 352 Direct Electron Transfer in Methane Oxidation 353 Challenges 353 Electron Transfer Across the Bridge Between Electrode and Cell 353 Mechanism of Cytoplasmic Reactions 354 Conclusion 354 Acknowledgment 354 References 355

15

Microbial Metabolism Kinetics and Interactions in Bioelectrosynthesis System 363 Zechong Guo and Chunxue Yang

15.1 15.2 15.2.1 15.2.2 15.2.3 15.3 15.3.1 15.3.2 15.4 15.4.1 15.4.2 15.4.3 15.5 15.5.1 15.5.2 15.5.3

Introduction 363 Microbial Metabolism Kinetics of Anode 366 Dynamic Description of Anodic Processes 366 Influencing Factors for Anodic Kinetics 371 Anode Potential Losses 372 Electrosynthesis Kinetics of Cathode 374 Dynamic Description of Cathode Electrosynthesis Processes 374 Cathode Potential Losses 376 Energy Balance in Bioelectrosynthesis Systems 378 Energy Transfer and Dissipation 378 Electron Balance in Electronic Circuit 380 Energy Recovery Efficiency Evaluation 383 Microbial Community Growth on Electrode 384 Anode Biofilm Formation Determined by Anode Potentials 384 Anodic Biofilm Structure 386 Interaction of Functional Communities in Integrated System of Bioelectrochemistry and Anaerobic Digestion 388 References 391 Index 395

xiii

Preface Bioelectrosynthesis: Here to Stay! Many things in science move in waves until they finally land. In the 1960s, with the advent of the fuel cell, it was found that microbial decomposition of organic matter could also be directly linked to power production in microbial fuel cells (MFC) [1]. The power production was modest, and the top ic lead a latent existence until the petroleum crisis of the late 1970s when novel routes for chemical production were explored. Visionary scientists then coupled this microbial electron transfer for the production of organic molecules such as glutamic acid and butanol [2–4]. At that time, technology did not follow and titers and rates were limited, not competitive to existing bioproduction approaches. Then, in the late 1990s, the old topic of the MFC resurfaced, with new discoveries on electron flow [5–7]. This time, technology had evolved as well, enabling higher production rates [8, 9] and also enabling different process outcomes such as hydrogen [10] or caustic soda [11] production. Unfortunately, for MFCs, even by then also, an alternative technology generating power from biomass had matured: anaerobic digestion. In this process, biogas is produced. Nowadays, anaerobic digesters can deal from small to large scale with complex waste streams, they are very robust, and most importantly they can deal with high loading rates. Top systems now convert over 50 kg organics/m3 reactor per day to methane. In terms of electron flow, this implies a current of almost 7000 A, going to methane. To my opinion, it will be extremely difficult for MFCs to become an alternative to this, certainly considering the higher complexity of the systems and the presently lower rate. No, besides the niches for MFC in sensing, the major promise lies in the return of the second wave: bioelectrosynthesis. Although there were some isolated reports on production of methane at cathodes in the late 1990s as well [6], around 2010, the topic truly resurfaced in the context of production of modest amounts of acetate from CO2 and electricity [12] in so-called microbial electrosynthesis (MES). Since then, titers and rates rapidly increased because of the availability of better technology, to now reach gram per liter levels [13]. Simultaneously, it is possible to also extract the product and thus obtain concentrates [14]. The product portfolio has expanded, from acetate to butyrate, caproate, and caprylate [15], toward alcohols such as ethanol [16] and isopropanol [17], even toward esters such as ethylacetate [18]. It appears that electricity-driven CO2 reduction

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is here to stay, and there are multiple good reasons for this: society is electrifying, which means that new applications are shifting to the use of electricity as energy source. The electricity is ubiquitously available, can be produced from renewable sources, and when used in the context of production leaves no traces such as salts in the water or the product. The coupling of electricity to CO2 conversion in the so-called carbon capture and utilization is rapidly emerging and MES will find its place within this portfolio. Already in 2010, we made the point that MES is more than reducing CO2 [19]. Many existing production processes are imbalanced in terms of electrons, requiring, e.g., the supply of very well-controlled amounts of oxygen, which complicates many production processes. Flynn et al. [20] showed elegantly that an anode could solve the electron imbalance, enabling production of ethanol from glycerol with an engineered Shewanella oneidensis strain. Later, Lai et al. [21] showed anode-associated conversion of glucose to α-ketogluconic acid at efficiencies over 90% by coupling the metabolism of Pseudomonas putida, remarkably a strict aerobe, to an anode. The use of this organism opens up an enormous array of novel production routes, multiple of the most attractive routes have recently been identified by Kracke and Krömer [22]. MES thus encompasses a broad range of production processes, both anodic and cathodic, both starting from CO2 and from substrate organics [23]. Similar processes emerge to produce methane or upgrade biogas and to produce inorganic products such as hydrogen peroxide or ammonia, many of which are discussed in detail in the following book and which all have the potential to evolve into mature technologies and processes. The challenges toward this are considerable and are both technological and microbial. When reading this book, grasp the excitement of this great field of science and engineering on the verge of breakthrough. This interface between biology and electrochemistry has already taught us many things about how microorganisms and microbial communities work, and they will continue to amaze us. Think about new, creative uses of bugs and electricity or how electron flow could affect our natural environment. Enjoy, Korneel Rabaey January 2018 Ghent University, Centre for Microbial Ecology and Technology (CMET), Faculty of Bioscience Engineering, Coupure Links 653, 9000 Gent, Belgium

References 1 Davis, J.B. and Yarbrough, H.F. (1962). Preliminary experiments on a micro-

bial fuel cell. Science 137 (3530): 615–616. 2 Hongo, M. and Iwahara, M. (1979). Electrochemical studies on fermentation.

1. Application of electro-energizing method to L-glutamic acid fermentation. Agri. Biol. Chem. 43 (10): 2075–2081.

Preface

3 Ghosh, B.K. and Zeikus, J.G. (1987). Electroenergization for control of H2

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transformation in acetone butanol fermentations. Abstr. Papers Am. Chem. Soc. 194: 79-MBTD. Kim, T.S. and Kim, B.H. (1988). Electron flow shift in Clostridium acetobutylicum by electrochemically introduced reducing equivalent. Biotechnol. Lett. 10 (2): 123–128. Kim, B.H., Kim, H.J., Hyun, M.S., and Park, D.H. (1999). Direct electrode reaction of Fe(III)-reducing bacterium, Shewanella putrefaciens. J. Microbiol. Biotechnol. 9 (2): 127–131. Park, D.H., Laivenieks, M., Guettler, M.V. et al. (1999). Microbial utilization of electrically reduced neutral red as the sole electron donor for growth and metabolite production. Appl. Environ. Microbiol. 65 (7): 2912–2917. Bond, D.R., Holmes, D.E., Tender, L.M., and Lovley, D.R. (2002). Electrode-reducing microorganisms that harvest energy from marine sediments. Science 295 (5554): 483–485. Rabaey, K., Lissens, G., Siciliano, S.D., and Verstraete, W. (2003). A microbial fuel cell capable of converting glucose to electricity at high rate and efficiency. Biotechnol. Lett. 25 (18): 1531–1535. Logan, B., Aelterman, P., Hamelers, B. et al. (2006). Microbial fuel cells: methodology and technology. Environ. Sci. Technol. 40 (17): 5181–5192. Liu, H., Grot, S., and Logan, B.E. (2005). Electrochemically assisted microbial production of hydrogen from acetate. Environ. Sci. Technol. 39 (11): 4317–4320. Rabaey, K., Bützer, S., Brown, S. et al. (2010). High current generation coupled to caustic production using a lamellar bioelectrochemical system. Environ. Sci. Technol. 44 (11): 4315–4321. Nevin, K.P., Woodard, T.L., Franks, A.E. et al. (2010). Microbial electrosynthesis: feeding microbes electricity to convert carbon dioxide and water to multicarbon extracellular organic compounds. mBio 1 (2): e00103–e00110. https://doi.org/10.1128/mBio.00103-10. Patil, S.A., Arends, J.B.A., Vanwonterghem, I. et al. (2015). Selective enrichment establishes a stable performing community for microbial electrosynthesis of acetate from CO2 . Environ. Sci. Technol. 49 (14): 8833–8843. Gildemyn, S., Verbeeck, K., Slabbinck, R. et al. (2015). Integrated production, extraction, and concentration of acetic acid from CO2 through microbial electrosynthesis. Environ. Sci. Technol. Lett. 2 (11): 325–328. Van Eerten-Jansen, M.C.A.A., Ter Heijne, A., Grootscholten, T.I.M. et al. (2013). Bioelectrochemical production of caproate and caprylate from acetate by mixed cultures. ACS Sustainable Chem. Eng. 1 (5): 513–518. Steinbusch, K.J.J., Hamelers, H.V.M., Schaap, J.D. et al. (2010). Bioelectrochemical ethanol production through mediated acetate reduction by mixed cultures. Environ. Sci. Technol. 44 (1): 513–517. Arends, J.B.A., Patil, S.A., Roume, H., Rabaey, K. (2017). Continuous long-term electricity-driven bioproduction of carboxylates and isopropanol from CO2 with a mixed microbial community. J. CO2 Util. 20: 141–149.

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18 Andersen, S.J., Berton, J., Naert, P. et al. (2016). Extraction and esterification

19 20

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of low-titer short-chain volatile fatty acids from anaerobic fermentation with ionic liquids. ChemSusChem 9 (16): 2059–2063. Rabaey, K. and Rozendal, R.A. (2010). Microbial electrosynthesis—revisiting the electrical route for microbial production. Nat. Rev. Microbiol. 8: 706–716. Flynn, J.M., Ross, D.E., Hunt, K.A. et al. (2010). Enabling unbalanced fermentations by using engineered electrode-interfaced bacteria. mBio 1 (5): e00190–10–e00190–17. Lai, B., Yu, S., Bernhardt, P.V. et al. (2016). Anoxic metabolism and biochemical production in Pseudomonas putida F1 driven by a bioelectrochemical system. Biotechnol. Biofuels 9. Kracke, F. and Krömer, J.O. (2014). Identifying target processes for microbial electrosynthesis by elementary mode analysis. BMC Bioinf. 15 (1): 410. Logan, B.E. and Rabaey, K. (2012). Conversion of wastes into bioelectricity and chemicals using microbial electrochemical technologies. Science 337 (6095): 686–690.

1

Section I Principle and Products Overview of Bioelectrosynthesis

3

1 Principle and Product Overview of Bioelectrosynthesis Fang Zhang 1 , Yuquan Wei 2 , and Guanghe Li 1 1 Tsinghua University, School of Environment and State Key Joint Laboratory of Environment Simulation and Pollution Control, Haidian District, Beijing, 100084, China 2 China Agricultural University, College of Resources and Environmental Sciences, Haidian District, Beijing 100193, China

1.1 Introduction With the pressing crisis of depletion of fossil fuels, the past decade has seen the significant growth in the use of renewable energy, which leads to the growing research efforts toward electricity production from solar, wind, wave, or biomass energy (as opposed to petroleum, coal, or gas) in a sustainable way [1]. As electricity produced based on these renewable sources is usually intermittent and off-grit, electrosynthesis has been considered as an effective strategy to store electrical energy from renewable sources in the forms of chemical compounds [2]. Adequate electrocatalysts are necessary to catalyze the electrode-driven chemical reactions, yet these chemical catalysts are usually too expensive to be scaled up for practical applications. As a result, biocatalysts, which can be an enzyme, an organelle, or even a whole cell, have drawn increasing attention in electrosynthetic processes because of their higher specificity and versatility [3]. Moreover, microbes as catalysts are inexpensive to grow and, if the microbes catalyzing the reactions gain enough energy for cell maintenance, are self-sustaining and long-lived. Therefore, bioelectrosynthesis represents a promising approach to store renewable energy or produce target chemicals in an energy-sustainable and low-cost way. Bioelectrosynthesis has emerged that electrical energy can be combined with biosynthesis to drive CO2 fixation as a means to directly produce the target compound or lead to the formation of acetyl-CoA and its derivatives for further synthesis. There are many assumptions for bioproduction in different pathways (Figure 1.1), which require inputs of solar energy or electrical energy (as an indirect solar derivative). One could speculate that instead of the Wood–Ljungdahl pathway (which would produce acetyl-CoA), the Calvin–Benson–Bassham cycle (which yields triose phosphates) can be driven on electrical current, leading to the formation of fermentable substrate from electricity and CO2 . This fermentable substrate could then further be used for bioproduction purposes. Bioelectrosynthesis: Principles and Technologies for Value-Added Products, First Edition. Edited by Aijie Wang, Wenzong Liu, Bo Zhang, and Weiwei Cai. © 2020 Wiley-VCH Verlag GmbH & Co. KGaA. Published 2020 by Wiley-VCH Verlag GmbH & Co. KGaA.

4

1 Principle and Product Overview of Bioelectrosynthesis

CO2

Power-driven Calvin cycle

Power driven Wood–Ljungdahl pathway

Light-driven Calvin cycle

Fermentable substrate

Acetyle-CoA and derivatives

Conventional fermentation Power-assisted fermentation

Cell-based synthesis

Target product

Cell-based synthesis

Figure 1.1 Overview of different routes toward bioproduction from CO2 . Source: Rabaey et al. 2011 [1]. Reproduced with permission of Elsevier.

Lastly, the fermentation itself can be complemented by electrical current to provide reducing equivalents to the cell. This can be considered as a hybrid metabolism when effective charge transfer occurs toward the cell. The major assumptions are summarized in Table 1.1. The theoretically achievable bioproduction densities for bioelectrosynthesis (product–carbon per hectare per annum) appear excessive at first glance. However, it is crucial to point out that photovoltaic panels are relatively efficient in capturing solar energy and that a first study producing acetate from CO2 has indicated high electron yields. Other factors such as CO2 and nutrient supply are likely to become limiting before these theoretical values are achieved. Therefore, electrosynthesis of organic compounds via abiotic or enzymatic catalysis of carbon dioxide reduction at electrode surfaces has been evaluated as a strategy for converting electricity into useful organic products for some time [7–9]. Bioelectrosynthesis relies on the use of biocatalysts on the electrode surfaces to achieve electricity-driven synthesis. For biocatalyst, it has the following advantages in the bioelectrosynthetic processes: (i) the high reaction specificity and controllability of enzymes and organelles, (ii) self-regeneration of the whole microorganisms as the catalyst, (iii) adaptation of the microbial (catalyst’s) quantity to the required conversion activity, (iv) flexibility in substrate use, (v) high versatility for product formation or conversion pathways, and (vi) improving the performance by decreasing the overpotentials at both anodes and cathodes [10–12]. However, microorganisms as biocatalysts are still far from perfect. Unlike true catalyst, microbes have been shown to consume part of the substrate or donor for growth albeit possibly only intermittently and are hard to keep a steady function or phenotype in different microenvironments. Microbial electrosynthesis (MES) is a form of microbial electrocatalysis, which is an emerging area in microbial electrochemical research and development. The concept of MES was used to describe the process when a microbial catalyst reduces CO2 into multicarbon chemical commodities with electrons derived

Table 1.1 Assumptions regarding the theoretical production rates as well as expected substrate requirements for bioelectrosynthesis, conventional fermentations, and algal bioproduction systems (c$ refers to dollar cent) [1, 4–6]. Aerobic fermentations

Anaerobic fermentations

Current-driven lithoautotrophy – aerobic

Current-driven lithoautotrophy – anaerobic

Algal production

C)a)

Glucose (c$0.6)

Glucose (c$0.6)

CO2 ($0)

CO2 ($0)

CO2 ($0)

Electron donor (cost c$/mol C)a)

Glucose (c$0.15)

Glucose (c$0.15)

Electricity (c$0.16)

Electricity (c$0.16)

Water (c$ 0)

Carbon source (cost c$/mol

Growth yield (mol C/mol C)

0.57

0.14

0.13

0.015

0.04–0.10

Maximal production density per hectareb)

AEM > CMM > CEM regarding the ability of preventing pH increase and the transport numbers for protons or hydroxyl ions [63]. 7.2.3

Operation Condition Optimization

After confirming the configuration of a microbial electrochemical cell for H2 O2 synthesis, such an operation condition is also needed to be further considered to improve the production yield and rate. Operation condition optimization was introduced by four aspects as follows. 7.2.3.1

Buffer Solution

As an important factor, pH could influence the overpotential in microbial electrochemical cell. It was reported that one unit increase for pH could lead to almost 60 mV decline of voltage based on the Nernst equation [85]; therefore, one cathode with a pH of 12 incurred approximately 300 mV of concentration overpotential [35] between the cathode and anode [86]. It was a complex process to control the cathodic pH because the OH− produced during H2 O2 synthesis. It was studied that adding CO2 to the cathode could reduce the pH and further reduce the concentration overpotential [87]. The added CO2 combined with OH− to form bicarbonate (HCO3 − ) and/or carbonate (CO3 − ), which had an effect on buffering the pH and decreasing the pH-related concentration overpotential [88]. It was testified that the difference exist in the pH of catholytes between nonbuffer and buffer solution with CO2 addition [89]. One test showed a decrease in cathode pH of 5 units roughly, along with reducing the applied voltage by 0.202 V with the addition of CO2 compared with the 0.3 V without CO2 [88].

7.2 Hydrogen Peroxide Produced in BES: Optimization and Application

The addition of CO2 formed the HCO3 − /CO3 and CO2 /HCO3 − couples, which had buffer action. Except for the bicarbonate buffer, phosphate buffer solution (PBS) with a regular concentration of 50 mM also played an important role in optimizing the process of H2 O2 production [42, 90]. 7.2.3.2

Hydraulic Retention Time

At a low hydraulic retention time (HRT) in continuous microbial electrochemical cell meant a high substrate concentration, which further effected on anode-respiring bacteria (ARB) kinetics. As one of the significant index for evaluating the system performance, the current density increased at low HRT implied an increased substrate utilization rate [91–93]. Sim et al. [60] compared the cumulative H2 O2 concentration and H2 O2 conversion efficiency for substrate in MEC reactor under different HRTs of 2, 6, and 10 hours. It generated a highest H2 O2 production rate of 144 mg H2 O2 /l/h at HRT of 2 hours, with a current density of 7.7 ± 0.6 A/m2 under a cathodic reaction time of 6 hours. However, it showed that the order of production rate was 6 hours>10 hours>2 hours under the cathodic reaction time of 24 hours. Also, when the reaction time increased to 24 hours from 6 hours, the H2 O2 production rate calculated by cathodic reaction time and the H2 O2 conversion efficiency decreased under all HRTs. It might be caused by self-decomposition of H2 O2 in liquid or reduction of H2 O2 to H2 O on the cathode [60]. A low HRT might get high H2 O2 conversion efficiency and H2 O2 production rate attributed to its high substrate concentration; however, decreased HRT might cause increased energy loss. It showed that energy loss became almost double from HRT of 2 to 10 hours. At low HRT, the process of hydrolysis and fermentation for complex organics into simple forms, which easily utilized by ARB [94], would be limited. It was necessary to find the appropriate HRT for high and stable H2 O2 production in future research studies. 7.2.3.3

Applied Voltage

It was first studied using an MFC reactor with an external call voltage of 0.5 V and could generate a higher yield of H2 O2 than those without energy input by Rozendal et al. [58]. As mentioned previously, such research studies had proved that adding an extra voltage could achieve the purpose of improving cell current density and further improving the yield and production rate of H2 O2 [35, 36, 95]. It was also an energy-saving process compared with the conventional electrochemical approaches because it needed a small external electrical energy input based on the direct provided energy from biological oxidation of substrate by electrochemical active bacteria at the anode [96, 97]. A high applied voltage could generate a relatively high current density, which was advantageous for anode performance in terms of organic degradation and coulombic efficiency. However, a relatively high voltage could also lead to the reduction of H2 O2 and the evolution of H2 [95]. Therefore, an appropriate applied voltage was necessarily combined with H2 O2 concentration, H2 O2 production efficiency at the cathode, and COD removal efficiency at the anode. Chen et al. studied the synthesis of H2 O2 in an MFC system at different applied voltages

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from 0.1 to 0.8 V to investigate an optimal voltage of 0.4 V and obtained a H2 O2 production rate of 2.12 kg/m3 /d and COD removal of 76% [95]. It displayed that the yield and rate of hydrogen peroxide production increased three to eight times compared with the MFC without energy input [95]. The H2 O2 concentration increased with the added applied voltage from 0.1 to 0.5 V and dropped off from 0.6 to 0.8 V [95]. It might be due to the fact that cathode potentials were close to the peak potential of two-electron oxygen reduction to hydrogen peroxide at an applied voltage of 0.1–0.3 V. With the increase of applied voltage to 0.5 V, the excess accumulation of H2 O2 increased and caused the slowed two-electron oxygen reduction rate. However, the production rate declined as the voltage increases when the applied voltage is >0.5 V in 8 hours. It might attribute to the H2 O2 reduction and H2 evolution stressed by the higher voltage simultaneously [95]. In another study, an optimized cathode was proven to have relatively stable performance to exhibit a high and stable H2 O2 production activity at an applied voltage of 0.6 V [41]. However, it was limited by the polarization of bioanode at a further increase of voltage to 0.7 and 0.8 V [41]. 7.2.4

Application of H2 O2 Production in BES

Fenton reaction as one of the oxidation process was an advanced technique for the treating recalcitrant organic contaminants by hydroxyl radicals (⋅ OH), which was generated from Fenton reagents (Fe2+ and H2 O2 ) [17]. Despite the high oxidative efficiency of Fenton reagents, the storage and shipment of concentrated H2 O2 limited the chemical Fenton application as well as the high cost of energy from electro-Fenton process [46]. Because the electrons supplied continuously from organics from wastewater [98–101], combined with Fenton reaction, the bioelectro-Fenton could drive the electro-Fenton process by using bioelectrons [102]. The bioelectro-Fenton had advantages such as the continuously H2 O2 produced avoiding the pressure of shipment and storage, easily conducted at ambient pressure and temperature, regenerated Fe2+ reducing the quantity of iron sludge, and enhanced mixing of reaction solution by oxygen or air sparging [46]. Bioelectro-Fenton offered an efficient and cost-effective approach for the removal of persistent organic pollutants from industrial wastewater including a high concentration aniline wastewater of 4460 mg/l [103]. Hydrogen peroxide production in BES combined with wetland water treatment was proven to generate high-quality and disinfected water without external input of chemicals [61]. In this combined system, suspended solids were filtered by the part of wetland, and then the soluble organics were fed into the anode chamber and supplied organic carbon to the microorganisms on the anode electrode. H2 O2 produced from the part of cathode chamber was used for disinfection of the wetland system [61]. The clean water from the wetland-BES system could be used for irrigation, even for direct human use instead of discharge to surface waters. The advantages of BES make the wetland operate during the whole year because the anode biocatalysts could adapt the low temperatures [104, 105].

7.3 Metal Ion Reduction in BES: Waste Treatment and Metal Reuse

7.2.5

Summary

Hydrogen peroxide produced by BES could be widely used in the disinfection of water or oxidation in many respects. Therefore, it was significant to improve the H2 O2 yield and production rate from optimizing electrode design, membrane selection, and operating conditions. Electrode usually applied the carbon-based material, through the adding surface area to make three-dimensional electrodes or combined with PTFE to constitute air-cathode whose surface layer with the ability of keeping the hydrophile–hydrophobe balance. For membranes used to separate two chambers and transfer ions, four common IEMs including CEM, AEM, BPM, and CMM were used in BES systems. Thereinto, AEM showed the best electrochemical performance, and then CEM > CMM > BPM. Moreover, regarding the ability of preventing pH increase and the transport numbers for protons or hydroxyl ions, the order was BPM > AEM > CMM > CEM. During the system operation period, buffer solution including bicarbonate buffer or PBS was used for reducing the pH and further reducing the concentration overpotential. Despite that the high applied voltage could generate a relatively high current density, a relatively high voltage could also lead to the reduction of H2 O2 and the evolution of H2 . A compatible applied voltage should be investigated to improve H2 O2 concentration, H2 O2 production efficiency, and COD removal efficiency simultaneously. In a word, to design a high continuous H2 O2 production system, and maintain its stable operation, these comprehensive factors should be discovered.

7.3 Metal Ion Reduction in BES: Waste Treatment and Metal Reuse Metal pollution was extremely harmful for environment; meanwhile, it was one kind of recyclable resource. It was estimated that 60% of copper waste was lost or landfilled directly [106]; the common treatment was incineration, which could reduce the volume of waste and generate some energy. The high concentrations of various metals from the remaining ashes could be recovered using solvent extraction and electrolysis [107, 108] after being leached out by chemical agents [109]. However, it was limited by high energy cost for electrolysis. Therefore, developing an economy technology with no sustainable way of secondary pollution for metal waste treatment and metal reuse was necessary. 7.3.1

Metal Waste Treatment

Hexavalent chromium (Cr(VI)) was widely used in industry production but was a priority toxic chemical, which was proven as a mutagen, teratogen, carcinogen, and highly corrosive [110]. Reducing Cr(VI) to Cr(III) by physical and chemical methodologies could reduce toxicity because Cr(III) has less toxicity, less solubility, and less mobility [111, 112]. However, these methods

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had disadvantages of high cost, as well as a chemical treatment could cause the secondary pollution by addition of reducing agents to wastewater. It was found that some aerobic and anaerobic microorganisms could reduce hexavalent chromium, but this microbiologically catalyzed technology needed the external carbon source addition [110, 113]. It was proven that the feasibility of reducing Cr(VI) in a biocathode MFC with graphite plate electrodes and Cr(VI) was removed as the form of Cr(OH)3 precipitate [110]. Some further research studies optimized the reduction rate of Cr(VI) and bioelectricity generation at the reactor design, electrode character, and system operation [114]. At a cathode to anode surface area ratio of 20, it gained the highest Cr(VI) reduction rate and power production of 20.6 mg/g VSS/h and 48 A/m3 , respectively, achieved in biocatalyzed graphite fiber cathode MFCs [114]. Hg2+ as another harmful metal ion was also proven to be reduced combined with electron electricity production in an MFC reactor [115]. Because it had high standard potential when accepting electrons, Hg2+ was also a kind of potential electron acceptor. Hg2+ was firstly reduced to Hg2 2+ in the solution of cathode chamber, then Hg2 2+ could be precipitated as the form of Hg2 Cl2 by a chemical reaction with the presence of Cl− principally in one research [115]. The result showed that both the initial pH and initial concentrations of Hg2+ could influence the removal efficiency of Hg2+ from electrochemical and chemical reactions. Because Hg2+ and Hg2 2+ were also further reduced to elemental mercury, the effluent mercury concentration could be reduced. Mercury existed as the dominant form of Hg2+ at pH < 3.0, and Hg(OH)2 at pH > 5.0 and both these two forms when pH value is between 3.0 and 5.0, as reported [116]. Wang et al. indicated that pH might influence the internal resistance and further effect on the maximum power densities, but no effect of H+ on the standard potential of Hg2+ or Hg2 2+ [115]. In addition, they found that the product was elemental Hg on the cathode surface and Hg2 Cl2 as deposits on the bottom of cathode chamber [115]. 7.3.2

Metal Reuse

At the cathode of BES system, several metal ions could be as electron acceptors and reduced from waste. Compared with traditional electrolytic technology for cathodic reduction of metal ions, a BES reactor had the advantage of just needing low or zero electrical energy input. The metal recovery at the cathode and wastewater treatment at the biological anode could be achieved simultaneously [24]. The high removal efficiency of pure copper crystals without CuO and Cu2 O enlarged the application range of MFCs [117]. However, this technology combined with copper recovery and electricity production was still as the early stage of development, and it might be limited by the decreased MFC performance caused by decreased copper concentration [117]. Different from Cu2+ , which could be reduced while controlling the cell voltage of zero without extra energy input, other metal ions including Pb2+ , Cd2+ , and Zn2+ needed applied voltages to recycle [24]. Modin et al. was the first to recover metals from a mixed solution containing Cu2+ , Pb2+ , Cd2+ , and Zn2+ [24] by using a bioelectrochemical reactor. It proved the feasibility of BES system in selective recovery of Cu, Pb, Cd, and Zn. Among these metal ions, Cu2+ was the easiest

7.4 Struvite Crystallization Recovery: Principle and Application in BES Systems

selective reduction ions by the cathode along with biological anode. When the applied voltage was approximately 0.34 V, Pb was recycled. Then, Cd was gained by adding up an extra voltage to 0.51 V, while Zn needed the highest voltage of 1.7 V [24]. Cu had a high recovery efficiency at the cathode of 77.2% because it could be deposited on the cathode surface as pure metals. The same with Cu and Zn also could be detectable as pure state without other metals contamination. On the contrary, small amounts of Cu could be detected when Pb was recovered, and Cd was recovered with Cu and Pb appearing in the deposits [24]. Cobalt was investigated as another metal that could be reduced from BES. The cobalt content in the lithium ion battery production might be higher than it was found in natural ores or even concentrated natural ores [118]. Commonly applied methods including pyrometallurgical and hydrometallurgical processes needed further dealing with adding alkaline chemical agents or optimized extractants with a high cost [119–123]. Also, the electrochemical processes applied in crystallizing Co2+ to cobalt metallic films needed extensive energy consumption [124, 125]. MEC supplied a way for recovery of cobalt and recycle of spent lithium ion batteries along with hydrogen generation. The process was optimized by studying the influence of different values including applied voltage, solution conductivity, pH, and temperature. It gained a yield of 0.81 mol Co/mol COD and 1.21–1.49 mol H2 /mol COD at applied voltages of 0.3–0.5 V [126]. 7.3.3

Summary

Metals existed in waste should be recovered and reused as the finite resources. BES offered a new thought to recover metals from contaminated streams. Compared with other technologies including physical, chemistry, and electrochemical method, BES was the most economical and friendly reduction method. Applying the BES technology to extract metals from contaminated wastewater or leachates made it possible with limited energy requirements while simultaneously treating wastewater. However, some factors including operating conditions and electrode materials might influence the reduction efficiency of metal ions or power generation in BES system. Therefore, optimizing this sustainable technology and gaining more recycle efficiency should be investigated in the future work. Also, the feasibility of this technology to other valuable metal recycling such as Ag needed to be investigated, proven, and optimized. Furthermore, the falling off of the metals deposited on the cathode and returning to the treated solution was also a problem to be considered [1].

7.4 Struvite Crystallization Recovery: Principle and Application in BES Systems Struvite (MgNH4 PO4 ⋅6H2 O) is a kind of white crystal, which is insoluble in water. It contains nitrogen and phosphorus nutrients without heavy metals and is regarded as a good slow release fertilizer. Crystallization is commonly formed in domestic sewage, animal wastewater, and sludge dewatering filtrate

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with nitrogen and phosphorus [127]. It is well known that phosphorus leaves the sewage system along with the excess sludge in the sewage treatment process. Therefore, the phosphate from the excess sludge was converted into magnesium ammonium phosphate precipitate by the struvite precipitation method. 7.4.1

Principle of Struvite Crystallization Recovery

Studies have shown that nutrients containing N and P from wastewater streams could be recovered as the form of slow-release fertilizers because of its low solubility and rich magnesium minerals [128]. Under certain conditions including supersaturation, pH, temperature, mixing energy, and the presence of foreign ions, struvite is formed at a same molar ratio of Mg, N, and P [29]. When the concentration of Mg2+ , NH4 + , and Hn PO4n 3− ions in the solution is greater than the solubility product constant of the struvite (K sp : 7.58 × 10−14 –4.36 × 10−13 ), precipitation will occur spontaneously [129]. The reaction equation is shown in equations as follows: Mg2+ + PO4 3− + NH4 + + 6H2 O → MgNH4 PO4 ⋅6H2 O

(7.4)

Mg2+ + HPO4 2− + NH4 + + 6H2 O → MgNH4 PO4 ⋅6H2 O + H+

(7.5)

Mg2+ + H2 PO4 − + NH4 + + 6H2 O → MgNH4 PO4 ⋅6H2 O + 2H+

(7.6)

The formation of struvite is a complex chemical precipitation reaction. The crystallization process is mainly divided into two stages: nucleation and growth. During the nucleation stage, the magnesium and ammonium ions react with the phosphate ions. Various ions form crystals and accumulate on the embryo to keep continuous growth and final formation of crystal during the growth period [130]. Most research studies focused on the effects of supersaturation ratio and pH (generally starting to crystallize at pH = 8) on crystallization, as these were most significant factors for struvite crystallization [129]. In recent years, researchers have studied to improve the recovered efficiency of struvite and some high value-added by-products from wastewater treatment processes by bioelectrochemical methods [131–133]. 7.4.2

Struvite Crystal Recovery Applied in MFC

In a BES system with a mixed solution containing phosphate, ammonium, and magnesium ions, the migration and enrichment of these ions were driven by the electrode potential and led to local supersaturation. Including the increased local pH, the precipitation of struvite crystals was natural. Phosphorus in the form of suspended solids is first dissolved in the process and then precipitated on the cathode to finally form crystals. It was proven that an MFC reactor could produce electricity and recover phosphorus from wastewater simultaneously in swine wastewater treatment by using a single-chamber MFC with air-cathode [25]. It showed a phosphorus removal of 70–82%. Solid deposits appeared while at the cathode surface in contact with the anode chamber electrolyte and indicated as struvite crystals by X-ray diffraction analysis.

7.4 Struvite Crystallization Recovery: Principle and Application in BES Systems

Recently, MFC has been widely used in struvite recovery, and many methods were investigated to improve its recovery efficiency. Adding the sea salt was one of the effective methods [131]. By adding brine to the system, the concentration of magnesium ions increased, thereby increasing the unsaturation and alkalinity required for the formation of struvite crystals [132]. 7.4.3

Struvite Crystal Recovery Applied in MEC

It was proposed that the single-chamber MEC system could produce hydrogen and synthesize struvite crystals to recover phosphorus simultaneously by Cusick and Logen [133] and developed a simultaneous production in microbial electrolysis–struvite sedimentation tank. The study improved the hydrogen production efficiency and the precipitation of struvite by using 304 stainless steel mesh as cathode, and the phosphorus removal rate reached 40%. At the same time, the hydrogen production consumed the energy input from part of recovered struvite that reduced the struvite. In addition, according to the principle of cathode struvite precipitation in MEC, Logan et al. studied the MEC reactor combined with the fluidized bed, which was more conducive to struvite crystallization and the simultaneous hydrogen production. The characteristics of hydrogen and phosphorus removal [134] further improved the output ratio of synthetic struvite crystals in MEC [135]. According to this principle, researchers have developed a dual-chamber MEC with a separation membrane [136]. By cleaning the cathode chamber and replacing the electrode after the end of each cycle, the struvite crystal could be recovered in time and cathode performance could restore, and this method could reduce the negative influence on the electrode performance caused by precipitate. Because of the consumption of protons in the cathode chamber, the entire cathode chamber was changed from neutral to alkaline, which was more conducive to the precipitation of struvite. It was shown that the precipitation rate of struvite crystal was positively correlated with the change of current and voltage in a certain range (0.4–1.2 V) [137]. 7.4.4

Summary

The primitive purpose for the formation of struvite in sewage treatment plants was the requirement of preventing scale formation through controlling the chemical dose. However, because struvite was as a good slow-release fertilizer containing nitrogen, phosphorus, and magnesium as well as satisfied the requirement of phosphorus removal, it could be used in wastewater treatment. At present, the recovery of struvite was still mainly achieved by adjusting the ion concentration and alkalinity, and the improvement of the BES system was based on carrying out these two aspects. More research studies on struvite recovery by BES systems should focus on improving the crystallization rate and recovery efficiency of struvite by improving the reactor construction and optimizing operating conditions.

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7.5 Ammonia Recovery and Other Inorganics Synthesis in BES Systems The main form of nitrogen in sewage is ionic NH4 + and NO3 − , which could migrate directionally driven by the internal electric field of BES system because of the charged characteristics of ionic substances. With the assistance of IEM, NH4 + and NO3 could be the removed and concentrated. BES system had the obvious advantages on removal of nitrogen-containing organic matter and ammonia recovery from the wastewater with high ammonia–nitrogen content. The process and influence factors of ammonia recovery in BES system were investigated as follows. 7.5.1

Migration of NH4 + in BES Systems

Ammonia recovery in BES was based on NH4 + passing through the IEM via migration or diffusion driven by current. Early studies have suggested that hydrogen ions moved from the anode to the cathode through a CEM (PEM) to balance the charge in MFC for generating electricity. However, more studies have found that the concentration of protons was much lower than other cations in anolytes. It was other cations moving through the IEM rather than protons. Based on this NH4 + loss mechanism, Cord-Ruwisch et al. studied the feasibility of pH control in MFC [26]. It showed that NH4 + accounted 90% of the ion flux in the system. The system was further developed as a gas exchange unit, in which the NH3 was recycled back to the anode for pH control. Kim et al. studied the NH4 + loss in single-chamber and dual-chamber MFC using swine wastewater, respectively, and found that it was enhanced by higher organic matter and N2 loading rates for NH4 + migration from the anode chamber to the cathode chamber [138]. Kuntke et al. also researched the relationship between NH4 + transfer rate and electricity generation under different NH4 + –N influent loads in a dual-chamber MFC with potassium ferricyanide as the cathode electron acceptor. The results showed the charge exchange stopped and the mobility of NH4 + remained stable while ion concentration was balanced from two electrode chambers [30]. Studies on the ammonia–nitrogen wastewater treatment by MEC have also confirmed the migration process of NH4 + driven by an applied voltage. The cathode cumulative amount of NH4 + was 318 mg/l, which was almost 10 times in an anode compartment because of the migration of NH4 + in the MEC [139]. 7.5.2

Ammonia Recovery in BES Systems

To further understand the migration mechanism of NH4 + , it was discovered that NH4 + was transported by two methods of migration and diffusion as well as the feasibility of ammonia recovery in BES. Thereinto, the diffusion of NH4 + was induced by the concentration gradient on both sides of the CEM. The migration of NH4 + was driven by the electric field. In BES system, NH4 + from the anode compartment was converted to NH3 (aq) by the high alkalinity produced in the cathode compartment. NH3 (aq) was further stripped by air or N2 for ammonia recovery, as well as absorbed by sulfuric acid. Therefore, ammonia can be

7.5 Ammonia Recovery and Other Inorganics Synthesis in BES Systems

e–

e–

(NH2)2CO + H2O → NH3 + CO2 NH3 + H2O → NH4+ + OH–

NH4+

NH4+ + H2O → NO2–/NO3– + H+ + e– NH4+ + NO2– → N2 + H2O

OH–

O2 + H2O + e– → OH–

NH4+ + NO2– → N2 + H2O

S2– + O2 → S/SO42– + e–

H

2+

Mg

SO42– + H+ + e– → S2– + H2O

Ca

Synthetic wastewater →

Na+

Mg2+ + Ca2+ + Na+ + K+ Anode chamber

NH4+ + O2 → NO2–/NO3– + H+

+

Organics → CO2 + NH3 + H+ + e–

Anode

NO2– NO3–

K+ CEM

2+

Mg2+ + NH4+ + PO43– + H2O →

SO42–

MgNH4PO4 · 6H2O

Cl–

Effluent of the anode chamber →

PO43–

SO42– + Cl– + PO43–

AEM

Cathode chamber

Cathode

Figure 7.5 Working mechanisms for resource recovery as well as wastewater treatment in a three-chamber MFC.

recovered directly in the form of NH3 in BES. The high pH value of the catholyte was one of the key factors for ammonia recovery, which could drive the ionic NH4 + to the free NH3 (aq). Therefore, ammonia could be recovered from some special wastewaters containing high concentrations of NH4 + –N (≥1000 mg/l) and low concentrations of organic compounds. Urine wastewater was considered to be suitable for ammonia recovery in BES because of its high electrical conductivity (20 ms/cm) and urea concentration (20 g/l). Kuntke et al. [30] treated urine wastewater for electricity generation and nitrogen removal simultaneously in an MFC reactor and successfully recovered electricity and a few grams of NH4 + per liter from high ammonia–nitrogen urine by means of volatilization. Lu et al. [140] developed a three-chamber resource recovery microbial fuel cell (RRMFC) for treating synthetic wastewater containing urine and various organic pollutants, as well as recovered nitrogen, phosphorus, and sulfur nutrients as shown in Figure 7.5. During the treatment, the hydrolysis of urea was increased by microorganisms and electrolysis. The self-generated electric field was used to drive ion migration, and then the nutrients were recovered from the wastewater. After one cycle (three days), the removal rates of urea, COD, histidine, creatinine, sodium acetate, SO4 2− , and PO4 3− were 99%, 97%, 99%, 91%, 99%, 98%, and 99%, respectively, as well as the recovered total nitrogen, PO4 3− , SO4 2− , and total salt of 42%, 37%, 59%, and 33% in the middle chamber. Arredondo et al. [141] conducted an economic analysis of ammonia recovered from urine wastewater in BES. Results showed that the energy required by ammonia recovery using BES was lower than other technologies including stripping and electrodialysis, which was proven feasible economically for urine treatment in BES system. 7.5.3

Other Inorganics Synthesis in BES Systems

The principle for application of BES was using the potential between the electrodes and the metabolism and catalysis of microbes, which could achieve the

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degradation of pollutants, the enrichment and precipitation of inorganic salts, as well as production of electricity, hydrogen, and methanogenesis. Because the interfacial effect and interfacial reaction of microbial membranes and electrodes in BES were relatively complex, it was easy to form a micro-oxygen microenvironment and changes in micropotentials within biofilms, which could create environmental conditions for the production of special compounds. Srikanth et al. [142] observed the formation of polyhydroxyalkanoate (PHA) in a microaerobic microenvironment on the biocathode of BES. PHA synthesis was as an alternative pathway to reoxidize nicotinamide adenine dinucleotide (NADH) under low dissolved oxygen levels. Electroactive biofilms (EABs) formed on bioanodes were discovered could produce polyhydroxybutyrate (PHB) or C3 and C4 compounds from amino acids and lactic acid in organic substrates [143]. Meanwhile, the EABs formed by pseudomonas could produce phenazine, which was a quorum-sensing biomolecule as an electronic shuttle from biofilm to anode [144]. Also, phenazine was discovered as an excellent antibacterial agent [143]. In addition, in the process of seawater desalination and brine desalination, HCl and NaOH could be compounded in the microbial desalting tank with IEM. A four-chamber microbial electrolysis desalination and chemical production pool was established, keeping the pH value of the anode chamber at around 7.0, which could greatly improve the microbial activity and generate the product of acid and alkaline shown in Figure 7.6. In the study, the addition of NaCl to the desalting compartment made the production of HCl and NaOH simultaneously in the microbial cells combined with desalination. Therefore, it achieved simultaneous desalination and chemical production [27, 28]. One-layered three-electrode BES equipment was reported to produce sodium hydroxide at the cathode, as well as current generation under the fixed anode potential, using sodium acetate as the electron donor for the anode [145]. It indicated the simultaneous sodium hydroxide synthesis and organic matter degradation in brewery wastewater treatment. Power

Resistance

NaOH

HCl

H+ + e–

+ + + + + +

− − − − − −

OH– H+

CO2

Organics

Anode

+ − + − BPM

+ + + + Cl–

+ + + +

− − −

Na+

− − − − −

AEM CEM

OH–

O2 H2 O

Cathode

Figure 7.6 The configuration for acid and alkaline production in a four-chamber microbial electrolysis desalination and chemical production cell.

7.6 Outlook

7.5.4

Summary

The principle and process of some inorganic synthesis including ammonia, PHA, PHB, phenazine, HCl, and NaOH in BES system were introduced as above. The ammonia recovery by BES system mainly utilized the migration of ammonium ions driven by internal electric field. The ionic state was converted into free state by the increase of enrichment alkalinity at the cathode and was finally recovered by steam stripping or acid absorption (sulfuric acid or boric acid). At present, the most research studies focused on utilizing BES system to recover ammonia from urine wastewater with high ammonia–nitrogen and low organic matter or other nitrogen-containing organic matter degradation wastewater. On this basis, including increasing the amount of electrode chamber and constructing an electrodialysis module stack were investigated in order to improve the performance and increase the yield of ammonia recovery. The adjustment of pH value in the system was significant for both ammonia recovery and simultaneous removal of nitrogen and phosphorus. More research studies regarding the critical control of alkalinity during simultaneous nitrogen and phosphorus removal could be further studied for improving resource recovery and treatment efficiency from wastewater.

7.6 Outlook In recent years, BESs have become a promising technology for wastewater treatment as well as energy and product recovery [1]. It could generate electricity power in an MFC system and biogas including hydrogen and methane in an MEC system [2–4]. Meanwhile, it was found feasible for some inorganic compound synthesis at the cathode in BESs [1]. One common product was hydrogen peroxide, which could be used in disinfection of water or oxidation in many respects. Compared with other traditional productive process, which always required highly toxic raw materials and solvents or high cost [18], generating H2 O2 by BES is a harmless and economical technology. Further research studies were regarding to improve the H2 O2 yield and production rate from optimizing electrodes design, membrane selection, and operating conditions [15, 63, 88]. Carbon-based electrodes were usually optimized by adding its surface area or combined with PTFE for H2 O2 generation in BESs. Metal ions could be electron acceptors and reduced at the cathode configuration. Copper recovery realized in an MFC without energy input, while other metal ions including Zn2+ and Pb2+ needed an extra voltage [24]. Other inorganic compounds such as struvite, ammonia, acid, and alkaline could also recover in BESs [25–28]. BES was valued for generating high value-added chemicals through cheap and sustainable resources and was expected to become a key process in the future development of biogeneration. Based on the principle of inorganic compound synthesis, an appropriate pH value, which could be adjusted by adding buffer solution (including bicarbonate buffer or PBS), as well as an appropriate applied voltage, should be further investigated for the optimization of recovery efficiency, organics removal, electricity, and biogas production in BESs. Furthermore, the

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falling off of the metals deposited on the cathode and returning to the treated solution again as well as the economic applicability also needed to be further considered.

Acknowledgments This research was supported by the National Natural Science Foundation of China (No. 51778607) and the Youth Innovation Promotion Association CAS (2017062).

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precipitation and liquid–liquid extraction techniques. Waste Manage. (Oxford) 31: 59–64. Jha, A.K., Jha, M.K., Kumari, A. et al. (2013). Selective separation and recovery of cobalt from leach liquor of discarded Li-ion batteries using thiophosphinic extractant. Sep. Purif. Technol. 104: 160–166. Krause, A., Uhlemann, M., Gebert, A., and Schultz, L. (2004). The effect of magnetic fields on the electrodeposition of cobalt. Electrochim. Acta 49: 4127–4134. Freitas, M. and Garcia, E. (2007). Electrochemical recycling of cobalt from cathodes of spent lithium-ion batteries. J. Power Sources 171: 953–959. Jiang, L., Huang, L., and Sun, Y. (2014). Recovery of flakey cobalt from aqueous Co(II) with simultaneous hydrogen production in microbial electrolysis cells. Int. J. Hydrogen Energy 39: 654–663. Wang, C.-C., Hao, X.-D., Guo, G.-S., and van Loosdrecht, M. (2010). Formation of pure struvite at neutral pH by electrochemical deposition. Chem. Eng. J. 159: 280–283. Lippens, C. and De Vrieze, J. (2019). Exploiting the unwanted: sulphate reduction enables phosphate recovery from energy-rich sludge during anaerobic digestion. Water Res 163: 114859. Le Corre, K.S., Valsami-Jones, E., Hobbs, P., and Parsons, S.A. (2009). Phosphorus recovery from wastewater by struvite crystallization: a review. Crit. Rev. Environ. Sci. Technol. 39: 433–477. Chen, Y., Liu, C., Guo, L. et al. (2018). Removal and recovery of phosphate anion as struvite from wastewater. Clean Technol. Environ. Policy 20: 2375–2380. Merino-Jimenez, I., Celorrio, V., Fermin, D.J. et al. (2017). Enhanced MFC power production and struvite recovery by the addition of sea salts to urine. Water Res. 109: 46–53. Ye, Z.-L., Chen, S.-H., Lu, M. et al. (2011). Recovering phosphorus as struvite from the digested swine wastewater with bittern as a magnesium source. Water Sci. Technol. 64: 334–340. Cusick, R.D. and Logan, B.E. (2012). Phosphate recovery as struvite within a single chamber microbial electrolysis cell. Bioresour. Technol. 107: 110–115. Ye, Z.-L., Deng, Y., Ye, X. et al. (2017). Application of image processing on struvite recovery from swine wastewater by using the fluidized bed. Water Sci. Technol. 77: 159–166. Cusick, R.D., Ullery, M.L., Dempsey, B.A., and Logan, B.E. (2014). Electrochemical struvite precipitation from digestate with a fluidized bed cathode microbial electrolysis cell. Water Res. 54: 297–306. Almatouq, A. and Babatunde, A. (2017). Concurrent hydrogen production and phosphorus recovery in dual chamber microbial electrolysis cell. Bioresour. Technol. 237: 193–203. Ye, Z.-L., Ghyselbrecht, K., Monballiu, A. et al. (2018). Fractionating magnesium ion from seawater for struvite recovery using electrodialysis with monovalent selective membranes. Chemosphere 210: 867–876.

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nia loss mechanisms in microbial fuel cells treating animal wastewater. Biotechnol. Bioeng. 99: 1120–1127. Villano, M., Scardala, S., Aulenta, F., and Majone, M. (2013). Carbon and nitrogen removal and enhanced methane production in a microbial electrolysis cell. Bioresour. Technol. 130: 366–371. Lu, S., Li, H., Tan, G. et al. (2019). Resource recovery microbial fuel cells for urine-containing wastewater treatment without external energy consumption. Chem. Eng. J. 373: 1072–1080. Arredondo, M.R., Kuntke, P., Jeremiasse, A. et al. (2015). Bioelectrochemical systems for nitrogen removal and recovery from wastewater. Environ. Sci. Water Res. Technol. 1: 22–33. Srikanth, S., Reddy, M.V., and Mohan, S.V. (2012). Microaerophilic microenvironment at biocathode enhances electrogenesis with simultaneous synthesis of polyhydroxyalkanoates (PHA) in bioelectrochemical system (BES). Bioresour. Technol. 125: 291–299. Pham, T.H., Aelterman, P., and Verstraete, W. (2009). Bioanode performance in bioelectrochemical systems: recent improvements and prospects. Trends Biotechnol. 27: 168–178. Rabaey, K., Boon, N., Höfte, M., and Verstraete, W. (2005). Microbial phenazine production enhances electron transfer in biofuel cells. Environ. Sci. Technol. 39: 3401–3408. Rabaey, K., Butzer, S., Brown, S. et al. (2010). High current generation coupled to caustic production using a lamellar bioelectrochemical system. Environ. Sci. Technol. 44: 4315–4321.

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8 Bioelectrochemical Ammonium Production – Nitrogen Removal and Recovery in BES Guoqiang Zhan Chinese Academy of Sciences (CAS), Key Laboratory of Environmental and Applied Microbiology, Chengdu Institute of Biology, No. 9, Section 4, Renmin Nan Road, 610041, Chengdu, Sichuan, China

Ammonium is a major pollutant in wastewater, which is classically removed by energy-intensive aerated nitrification and then via heterotrophic denitrification with additional carbon resources. The aerobic nitrification process can take up 60% of the total energy consumption of a wastewater treatment plant for nitrogen removal [1]. It is increasingly approved that a sustainable approach should be innovated in nitrogen removal aiming at energy efficiency and resource recovery from wastewater. Recently, ammonium recovery from electrochemical or bioelectrochemical systems (BESs) has been proposed as an alternative to traditional ammonium removal or recovery approaches [2–4]. However, bioelectrochemical nitrogen removal/recovery from wastewater is more energy efficient compared to electrochemical strategy, as microbes at the anode can convert chemical energy stored in substrates into electric energy [5–7]. BESs are a novel treatment technology to remove nitrogen from wastewater coupling to electricity generation in a microbial fuel cell (MFC) or hydrogen production in a microbial electrolysis cell (MEC), in which different types of nitrogen such as ammonium and nitrate have been addressed either at the anode or at the cathode in the field of nutrient removal and recovery [8–12]. Figure 8.1 shows the nitrogen cycle in BESs. Alternatively, ammonium can be recovered from the wastewater by stripping under alkaline conditions and then the gaseous NH3 can be absorbed by acid for ammonium sulfate and ammonium carbonate production. Addition of substantial amounts of chemicals is needed for maintaining desirable alkaline conditions for forming both struvite and gaseous NH3 . Until now, BESs have gained popularity for removal and recovery of nitrogen from wastewater. Recent excellent reviews highlight the potential application of BES for recovering nutrients and energy while treating wastewater [2, 15]. The objective of this review is to give a summarization of the recent advances on nitrogen transformation in BESs, including ammonium migration and recovery, anodic ammonium oxidation, and nitrification/denitrification, as well as the existing problems and challenges. Bioelectrosynthesis: Principles and Technologies for Value-Added Products, First Edition. Edited by Aijie Wang, Wenzong Liu, Bo Zhang, and Weiwei Cai. © 2020 Wiley-VCH Verlag GmbH & Co. KGaA. Published 2020 by Wiley-VCH Verlag GmbH & Co. KGaA.

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MEC MFC

e c NH3

NH3

H+

j

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d NH4+

b

f

CH2O0.5N0.2

NH4+

Biomass

Anode

i –

h

NO3

CEM / AEM

H2O/O2 g

N2

N2 Cathode

Figure 8.1 Schematic diagram showing possible nitrogen cycle in bioelectrochemical systems [3, 13, 14]. (a) Anaerobic ammonium oxidation with electrode as the sole electron acceptor, (b) ammonium nitrogen in cell synthesis, (c) ammonia transport through diffusion, (d) ammonium ion transport through CEM by electricity-driven migration, (e) ammonia stripping, (f ) nitrification, (g) denitrification, (h) ANAMMOX, (i) hydrogen evolution or oxygen reduction, and (j) conversion of ammonium to ammonia under alkaline pH.

8.1 Ammonium Migration and Recovery In a two-chambered MFC or MEC, the anode and cathode are separated by a cation (or proton) exchange membrane (CEM). Ammonium as a cation can permeate CEM and is used as a medium for transporting protons from anode to cathode. Here, ammonium migration and ammonia diffusion across the CEM are the two principal processes for ammonium transfer [16]. Migration of ammonium (NH+4 ) is driven by the electric field, while the diffusion of ammonia (NH3 ) is caused by the concentration gradient across the membrane [17]. Ammonium migrates via the membrane and accumulates in the cathode to allow charge balance in the double-compartment system. The ammonium ions in the cathode can be volatilized as ammonia gas via gas stripping in alkaline pH and then absorbed by dilute acid to form ammonium sulfate or carbonate [17]. Alternatively, ammonium is deposited as struvite (NH4 MgPO4 ) with stoichiometric amounts of phosphorus and magnesium addition [7]. However, the unbalance ratio of ammonium, magnesium, and phosphorus in wastewater make this approach difficult in practical application, especially additional alkali and magnesium. Ammonium recovery from swine wastewater has been studied by Kim et al. in single- and double-compartment MFCs [18]. However, the removal efficiency

8.1 Ammonium Migration and Recovery

of ammonium is only 60% with a long time of five days of start-up in a single-compartment MFC. The removal pathway is ammonia migration and diffusion. The ammonium migration can be enhanced by shifting BES from MFC to MEC. In MEC, hydrogen and alkali are generated rapidly at the cathode and act as the driving force for volatilizing ammonia. Wu and Modin recover ammonium from reject water combined with hydrogen generation [19]. In the anode, bacteria convert organic matters into carbon dioxide and release electron to the external circuit and then for hydrogen generation at the cathode with 96% efficiency. In this study, the anode was fed with synthetic wastewater, while the cathode compartment received either synthetic or real reject water with 1 g NH+4 –N/l. The gas (hydrogen and ammonia) evolved from the cathode compartment was passed through an acid adsorbent solution for recovering ammonium. Air stripping of the catholyte was necessary for volatilization of ammonia from the reject water. Ammonium recovery efficiencies reach to 94% with synthetic wastewater and 79% with real wastewater. This process is important to demonstrate electron-flow-driven migration of ammonium against a concentration gradient [19]. Traditional ammonia stripping processes need aeration and alkali addition and then acid adsorption, which require 32.5 kJ g/l ammonium-N [7]. In contrast, according to the calculation, ammonium recovery in BES can produce a net energy yield of 3.46 kJ g/l ammonium-N [6]. Therefore, ammonium recovery by BES operation is significantly energy efficient, which is less energy requiring compared to that by ammonia stripping. Urine and landfill leachate contain a large amount of nutrition, which is a potential resource for recovering ammonium. In domestic wastewater, 80% of nitrogen comes from urine [7]. Therefore, a source separation of urine is beneficial for nitrogen recovery and decreasing nitrogen removal load on wastewater treatment. Besides, the conductivity of urine is about 20 ms/cm with the urea concentration of about 20 g/l and a high buffering capacity [1], which is a suitable feed for BES. Ammonium recovery from diluted urine was studied in MFCs [20]. Later, the recovery concept was pointed out for source-separated urine and showed an ammonium recovery rate of 3.3 g N/m2 /d with a power output of 250 mW/m2 in MFCs by Kuntke [6]. Ammonium was transported from anode into cathode in a two-chamber cell via the mechanisms of ammonium migration and ammonia diffusion across membrane. In the cathode, ammonium can be transferred into ammonia under alkali condition and then recovered by volatilizing and absorbing into a dilute sulfuric acid for ammonium sulfate production. Subsequently, they also studied ammonium recovery from urine in MEC. A high recovery rate of 162 g NH4 + –N/m2 /d was obtained with a power requirement of 8.2 MJ/kg NH4 + –N, coupling to hydrogen generation, which can be used as the drive force for air stripping [21]. Recently, a study showed an economic analysis of ammonium recovery from urine through BES operation, indicating that the application of this approach for urine treatment was less energy intensive compared to other ammonium recovery technologies such as stripping, freeze–thawing, and electrodialysis [10, 22]. Landfill leachate solutions are also a potential feed for BES because of the high conductivity, high buffering capacity, and high ammonium.

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To simultaneously recover ammonia and generate energy, MECs are a better option because hydrogen production in MECs usually occurs under high current condition. However, without aeration in the cathode chamber of an MEC, ammonia stripping out of the catholyte could be challenging. It is found that there is little ammonia recovery in the absence of aeration and the catholyte accumulates a high concentration of ammonium at 8202 mg/l [15]. Therefore, supplying aeration was shown to significantly drive ammonia out of the cathode and resulted in recovery of ammonium at 0.77 M in the recovery bottle [23]. Aeration could increase the energy consumption of BES, and thus development of an appropriate method to drive ammonia out of the cathode without significant energy input will be important to implement ammonium recovery in BES. The opportunities and challenges for resource recovery from leachate via BES operation were reviewed, as well as the perspectives for future development [15].

8.2 Anodic Ammonium Oxidation In BESs, the anode can serve as an alternative electron acceptor and substitute oxygen for substrate oxidation such as organic matters, sulfide, and ammonium [24]. In fact, anaerobic ammonium oxidation is demonstrated in ANAMMOX process by bacteria, which use nitrite as the electron acceptor for ammonium oxidizing to mainly nitrogen gas and small nitrate [25, 26]. In addition, the finding of natural Fe–ANAMMOX and Mn–ANAMMOX in soil further supports the potential of ammonium oxidation via electrode as an acceptor because the microbial metal reduction shared the similar mechanism with anodic oxidation of BES on the extracellular electron transfer [27–29]. Theoretically, ammonium oxidation exhibits negative Gibbs free energy (−357 kJ/mol) for anaerobic oxidation under standard conditions, indicating that ammonium can be used as a substrate for electricity generation in a BES with ammonium as an electron donor in anode and nitrite as an electron acceptor in cathode [30]. Until now, several studies have investigated the anodic ammonium oxidation in MFC or MEC [8, 13, 14, 31–33]. In this case, the anode can substitute oxygen and saved the energy cost for aeration nitrification. In one study, an MFC used to treat swine wastewater showed a high removal performance of ammonium, but the mechanistic details of ammonium oxidation were not studied [18]. He et al. reported ammonium-dependent generation of electricity in a MFC, indicating that the addition of ammonium to organic-free MFCs could generate electricity with a low coulombic efficiency (CE) of 0.06–0.34% [8]. Besides, nitrifying bacteria and denitrifying bacteria were present at both of the electrodes, whereas no ANAMMOX bacteria were explored. The authors concluded that the dominant ammonia oxidizing bacteria (AOB) were closely related to Nitrosomonas europaea. Zhan et al. revealed that anodic ammonium oxidation was coupled to cathodic denitrification in a single-compartment three-dimensional MEC via applying low voltage

8.2 Anodic Ammonium Oxidation

from 0.2 to 0.4 V [33]. The following stoichiometric equations described the corresponding anodic and cathodic reactions: NH+4 + 2H2 O → NO−2 + 6e− + 8H+ (anode) 2NO−2 + 6e− + 8H+ → N2 ↑ +4H2 O(cathode) Oxygen was supplied as electron acceptor for nitrite production in traditional nitrification because the oxidization of 1 mol ammonium coupled to 2 mol nitrite reduction according to anodic and cathodic reactions. It was indicated that a part of ammonium was first converted to nitrite under microaerobic conditions by autotrophic nitrification, which then served as the cathodic electron acceptors coupling to anodic ammonium oxidation. The nitrogen removal efficiency increased from 70.3% to 92.6% with the increase of applied voltage from 0.2 to 0.4 V, and the maximum current varied from 4.4 to 14 mA, indicating that the increasing applied voltage can enhance autotrophic nitrogen removal [33]. The possibility of ammonium oxidation with anode as the sole electron acceptor was also demonstrated in a dual-chamber MEC with ammonium as the sole electron donor [13]. Approximately 95% of ammonium was oxidized into nitrate, as well as small nitrite [13]. N. europaea (40.3%) and the genus Empedobacter were the dominated microbial community on the anode. Soluble microbial products (SMPs) were detected, which enhanced electron transfer and current generation [13]. Previous studies pointed out that Pseudomonas sp. can enable other bacteria to achieve extracellular electron transfer via producing metabolites as electron shuttles. The authors declared that a similar strategy was used by the genus Empedobacter to enable N. europaea to achieve extracellular electron transfer for anodic ammonium oxidation under anaerobic conditions [13]. With the microaerobic conditions, the mechanism and pathway of ammonium oxidization was revealed in a MFC. In fact, nitrite can be oxidized by electrochemical reaction for current generation. Chen et al. revealed that ammonium was oxidized firstly to nitrite by AOB, and then the nitrite constituted a chemical cell to power generation [31]. Besides, they also developed an AO-MFC and declared that the electron flowed from ammonium to ammonium monooxygenase, Cyt aa3 oxidase, and anode, which were used for triggering ammonium oxidation, ATP synthesis, and current generation. Dissolved oxygen played a critical role in electron distribution [32]. DO concentration too high (more than 6.45 mg/l) or too low (less than 0.5 mg/l) could have great negative effects on the performance of electricity generation and total nitrogen removal [32]. In addition, with the strictly anaerobic conditions, ammonium can also be oxidized into nitrite or nitrogen gas in BES via controlling different electrode potentials. Two papers reported that ammonium can be converted into nitrogen gas using microbes as biocatalyst on the anode with positive electrode as electron acceptor [13, 14]. Microbial community analysis of 16S rRNA genes showed a diverse community that integrated members of varied functional bacteria involved in nitrogen transfer. Zhan et.al revealed that the dominant genera of Nitrosomonas, Comamonas, and Paracocus could be important for the

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16S rDNA sequencing

Fluorescence in situ hybridization (FISH) Brocardia and kuenenia genera

Actinobacteria (3%) Hydrogenedentes (1%) Firmicutes (2%) Bacteroidetes (5%) Proteobacteria (74%) Planctomycetes (7%) Unclassified bacteria (8%)

50 µm

Nitrosomonas and Nitrosospira genera

Alphaproteobacteria (44%) Betaproteobacteria (26%) Deltaproteobacteria (1%) Gammaproteobacteria (3%) 50 µm

Figure 8.2 16S rDNA analysis of the biofilm in one representative nitrifying BES anode compartment. On the left, phyla identified based on sequencing are shown. On the right, representative images of reactor samples after fluorescence in situ hybridization (FISH) are shown using specific fluorescent probes for anoxic ammonium-oxidizing (ANAMMOX) bacteria such as Brocardia and Kuenenia and AOB as Nitrosomonas and Nitrosospira. Source: Vilajeliu-Pons et al. 2018 [34]. Reproduced with permission of Elsevier.

electron transfer from ammonia oxidation to anode [14]. Vilajeliu-Pons et al. showed that nitrifying bacteria (Nitrosomonas genus), ANAMMOX bacteria (Brocardia and Kuenenia genera), denitrifying bacteria (several Bacteroidetes and Proteobacteria members), Feammox (Actinobacteria members), and Firmicutes were identified inside the anode compartment [34]. The entire diversity of bacteria was not only involved in the ammonium oxidation but also contributed to organic matter degradation (Figure 8.2). This result is in accordance with the CE of 35%. The high bacterial diversity is in agreement with the previous conclusion reported by Qu and Zhan and coworkers, where N. europaea was identified as the major contributor to a community together with Empedobacter [13], or a minor genus in a community dominated by the denitrifying Thermomonas [35]. As ammonium can be oxidized into nitrite anaerobically in BES, the nitrite production compensated the lack in the feed to accelerate the ANAMMOX process. A single-compartment BES with an anodic potential of −500 mV (vs. Ag/AgCl) realized ammonium oxidization to enhance ANAMMOX without nitrite addition [36]. The AOB bacteria coexisted with ANAMMOX bacteria in the presence of the anodic potential. The improvement of ANAMMOX was further confirmed while increasing the influent ammonium concentration [36]. However, the balance of ammonium oxidization and nitrite reduction was disturbed by the insufficient cathodic electron acceptors such as nitrite and nitrate.

8.2 Anodic Ammonium Oxidation

Electrode potential controlling is an advisable strategy for adjusting chemical reactions such as cathodic hydrogen, methanogenesis, and acetic acid production [37]. Therefore, anode potential can also be controlled by applying DC-POWER or potentiostat to realizing different production of anodic ammonium oxidation. In MFC, the dynamic process of ammonium oxidation is very slow, while the reaction rate can be improved in MEC, as well the CE. In a study, the electrode potentials of ammonium oxidation into nitrite and nitrite are −0.34 V and −0.77 V, respectively. Therefore, the authors controlled the anodic potential at −0.5 V for nitrite production rather than nitrate [36]. Several different anodic potentials are also reported for ammonium oxidation. A polarized electrode of +0.6 V (vs. Ag/AgCl) was used for ammonium mainly oxidizing into nitrogen gas, and the maximum rate of ammonium removal reached 60 mg/l/d [14]. A special oxidation peak occurred at 0.6 V, showing that the bioanode possess good electroactivity for ammonium oxidation. Vilajeliu-Pons et al. pointed out a complete anoxic ammonium oxidation to nitrogen gas with the anode potential fixed at 0.4 V (vs. Ag/AgCl) concomitant with little production of nitrite, nitrate, and nitrous oxide [34]. They also showed an oxidative potential of 0.43 V, indicating a thermodynamically available anodic ammonium oxidation. The removal rate of ammonium was comparable to traditional WWTP with 35 ± 10 g N /m3 /d [34]. An anaerobic cathode was also maintained at a potential equal to that of an aerobic chamber in order to eliminate the effects of oxygen diffusion from cathode to anode. With imposed potential of positive 67 mV to the cathode, the ammonium removal efficiency was obtained at 12.67%, which was higher than that of 9.3% at an aerobic cathode [38]. The authors declared that the improved ammonium removal efficiency with potential control was due to the better ANAMMOX bacteria activity in the anodic compartment during strictly anaerobic conditions. In their study, ammonium ions was firstly oxidized into nitrate instead of direct nitrite production (Eq. (8.1)), and then nitrate was converted into nitrite (Eq. (8.2)), which was used as the acceptor for the remaining ammonium removal in the ANAMMOX process [38]. NH+4 + 3H2 O → NO−3 + 8e− + 10H+

(8.1)

NO−3 + 2H+ + 2e− → NO−2 + H2 O

(8.2)

In the anode oxidation process, CE is a critical parameter for evaluating the performance of anodic ammonium oxidation. In the mixed system of ammonium and organic matters, the traditional calculation of CE is not accurate because of the electron flux from the organic matters oxidation. Nitrite is also an electrode donor for current generation. Chen et al. constructed a novel inorganic MFC based on anodic nitrification [31]. The authors indicated that several inorganic substrates apart from ammonium, and the intermediates such as nitrite and hydroxylamine could also serve as fuel substrates for current generation. However, the CE in their study was 0.31–1.1% at a very low level [31]. Until now, the reported research on anodic ammonium oxidation in BES showed a very low CE except a three-dimensional electrode designed by Zhan et al. [33]. In fact, the CE is direct related to the current density, which is dependent on many transport processes, including the fluxes of electron donor, electron acceptor, ion, acidity,

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and alkalinity at anode and cathode, respectively. Several design approaches are used for improving the function of BES. A high electrode surface area is beneficial for a high volumetric current density, such as a 3D electrode structure. Besides, membrane/electrode assemblies (MEAs) where a porous anode is in close contact with the membrane can obtain a high current density because of the short distance between anode and cathode. However, the close distance will lead to oxygen diffusion from cathode to anode, resulting in a low CE. Therefore, in order to prevent crossover from hindering current production, high concentrations of substrate are often needed. The critical transport rates that limit the performance of microbial electrochemistry technologies (METs) has been reviewed in more detail on the transport processes by Popat and Torres [39].

8.3 Nitrification/Denitrification in BESs Nitrate and nitrite can be used as electron acceptor for autotrophic denitrification at the biocathode, using the electrons for the reduction reaction. A wide diversity of microbes can electrochemically interact with electrodes, such as Shewanella, Pseudomonas, and Geobacter species [40]. Microorganisms in the family Geobacteraceae were found to utilize electrodes as electron donors for nitrate reduction [41]. Gregory et al. firstly demonstrated that nitrate can be reduced into nitrite with a graphite electrode as the sole electron donor [42]. They reported that the maximum hydrogen production rate at the cathode was 6 × 10−7 mmol/h, which was 10 000-fold slower than the rate of current consumption by the nitrate reducing bacteria [42]. Therefore, the drive force of nitrate reduction directly comes from the cathode rather than hydrogen autotrophic denitrification. Clauwaert et al. showed that nitrate was removed at a high rate of 146 g N/m3 /d and the complete denitrification could be realized in a tubular dual-chamber MFC [43]. In this device, acetate was used as the electron donor in the bioanode, while nitrate was used as the electron acceptor in the biocathode. The process of acetate oxidation can release electron to the anode and then deliver to the external circuit for electricity generation coupling to the cathodic nitrate reduction into nitrogen gas. Nitrate is easily reduced at cathode by bacteria. The potential and applied voltages are very important for the product. Su et al. reported that the initial concentration of nitrate had effects on the products of reduction [44]. With the increase of nitrate, the rate of denitrification was increased at the cathode potential of negative 0.5 V (vs. Ag/AgCl) but dissimilatory nitrate reduction to ammonium (DNRA). While nitrite is used as the acceptor, the electron recovery via denitrification and DNRA were 50.51% and 43.06%, respectively [44]. Zhang et al. also pointed out that the more negative potential was inclined to ammonium formation instead of nitrogen gas. When the cathode potential was set from −0.3 to −1.1 V (vs. Ag/AgCl) at an initial nitrate concentration of 100 mg NO−3 –N/l, the DNRA electron recovery increased from 10.76% to 35.06%, while the denitrification electron recovery decreased from 63.42% to 44.33% [45]. In addition, electrolysis of water can provide oxygen at the anode and hydrogen at the cathode, which may consume more electric energy than MFC or MEC.

8.3 Nitrification/Denitrification in BESs

Mellor firstly reported nitrate and nitrite reduction via immobilizing nitrate and nitrite reductase on the cathode [46]. Sakakibara and Kuroda also immobilized denitrifying bacteria on the cathode for nitrate reduction [47]. In this design, the rate of denitrification is linearly related to the current density. Islam and Suidan designed an electrode biofilm reactor with the applied electric current intensity varied from 0 to 100 mA, and the nitrogen removal efficiency reached to 98% at a current of 20 mA when phosphate was used as a buffer [48]. Sakakibara used carbon rod as the anode, which can offer carbon dioxide maintain the pH at 7–8 [49]. The electrode biofilm reactor with carbon rod can release carbon resources, oxygen, and hydrogen for microorganism’s growth and nitrate removal, but the low conductivity of underground water and the low growth rate of autotrophic bacteria limits the application of electrode biofilm reactor for nitrate removal from underground water. Some improvements are pointed out via three-dimensional electrodes, which also called particle electrode or bed electrode, improving the electrode surface and current efficiency. Watanabe and coworkers constructed an electrode biofilm reactor, which used the oxygen evolution for nitrification at anode and then the anode electrolyte flowed into the cathode chamber for nitrogen removal via hydrogen autotrophic denitrification [50]. In this device, oxygen and hydrogen will be used sufficiently for nitrification and denitrification, respectively. The 3-D electrode improved the nitrate removal efficiency, and a high removal load can reach to 0.288 mg NO−3 –N/cm2 /d reported by Zhou et al. with an applied current [51, 52]. In fact, ammonium is the major nitrogen resource in wastewater. Therefore, nitrification is necessary before denitrification. Nitrification and denitrification performed at the cathode were useful and efficient for ammonium removal via BES. Some wastewater including urine and landfill leachate contain high amount of ammonium, which should be considered for nitrogen recovery rather than removal. Recently, too many papers reported nitrogen removal in BES via nitrification and denitrification at the cathode for treating some low concentration of ammonium such as domestic wastewater [2]. Virdis et al. constructed an MFC with an addition nitrification reactor for nitrogen removal from ammonium. While the ammonium ions can penetrate through CEM from anode to cathode, leading to the effluent with ammonium residual [53] (Figure 8.3a). Later, they designed a cathodic biofilm for simultaneous nitrification and denitrification via dissolve oxygen control [12] (Figure 8.3b). The results showed that nitrifying bacteria occurred on the outer biofilm and denitrifying bacteria presented inside the biofilm, offering a true and correct explanation for ammonium removal process. The nitrogen removal load was 0.41 kg /m/d with a maximum power density of 34.6 W/m3 [12, 56]. Dissolved oxygen is a critical factor for simultaneous nitrification and denitrification because the low level of oxygen does not facilitate ammonium oxidization and the high level of oxygen does not facilitate denitrification. Therefore, an independent nitrification reactor before cathode denitrification is necessary for completing ammonium removal. Hence, a double MFC with aerobic cathode and anaerobic cathode was performed by supplying the anode effluent for aerobic cathode influent and successively for anaerobic cathode [54] (Figure 8.3c). The double-MFC device can be expanded to 50 l or even more with stable ammonium oxidization and nitrogen removal

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Figure 8.3 The BES systems designed for complete nitrogen removal involving nitrification and bioelectrochemical denitrification. (a) An MFC with an external nitrifying reactor. Source: Virdis et al. 2010 [12]. Reproduced with permission of Elsevier. (b) Simultaneous nitrification and denitrification in the cathode of an MFC. Source: Virdis et al. 2008 [53]. Reproduced with permission of Elsevier. (c) Two MFCs with aerobic and anaerobic cathodes, respectively. Source: Xie et al. 2011 [54]. Reproduced with permission of Elsevier. (d) A tubular MFC with double cathodes. Zhang and He 2012 [55]. Reproduced with permission of Elsevier.

rate. Zhang and He [55] designed a single-anode device with a double-cathode device, which used CEM separating the anode and aerobic cathode, while anion exchange membrane separating the anode and anaerobic cathode. Both batch and continuous experiments were carried out for determining nitrogen removal performance (Figure 8.3d). The results showed that partial heterotrophic denitrification enhanced total removal efficiency. The authors used this device for digested sludge and landfill leachate treatment, and the COD removal efficiency reached to 80% and the total nitrogen removal efficiency reached to 50–70% with a low energy consumption of 0.06 kWh/m3 or 0.1 kWh/kg COD [57]. In order to cut down the energy input, the air cathode is considered as an efficient operation without air pump. Zhang et al. reported a positive result via air cathode MFC for treating domestic wastewater and the nitrogen removal efficiency reached to 76% [58].

References

8.4 Existing Problems and Challenges BES has gained popularity for removal and recovery of nitrogen from wastewater, but there are still several challenges to overcome before the technology can be applied in practice. Nitrogen removal in the anode or cathode is dependent on some critical factors including (i) substrate concentration, (ii) ammonium coexist with organic matters or other ions, (iii) current density and applied voltage, (iv) the diversity of microorganisms, and (v) the conductivity and pH of electrolyte. Many of these factors are interrelated with each other, e.g. the substrate concentration and conductivity and current density. Therefore, the interactions should be studied in more detail to improve the nitrogen removal and recovery efficiency from real wastewater such as domestic wastewater, feces, urine, and landfill leachate. In order to understand the nitrogen transformation processes, which determine ammonium transport, modeling the system will give valuable insights and understanding of its performance for application. Urine contains about 10 g/l of COD and 9 g/l of ammonium. Ideally, 0.57 kg of COD is required for the removal of 1 kg N. Therefore, urine has sufficient electron donor for ammonium recovery without additional carbon. However, in this ideal system, it is assumed that all the COD is converted into electrons and ammonium is the single transported cation. Therefore, in real wastewater, the efficiency of COD and nitrogen removal, CE, and cathode efficiency can be obtained in BES. Low-energy demand and energy generation are attractive for waste treatment. However, the treatment efficiency is not high while using BES system only. Therefore, BES should be combined with other techniques for different aims (energy recover or waste removal). Presently, nitrogen removal and recovery, and COD removal, as well as phosphorus compound recovery, will be more valuable than electricity generation. Although ammonium can be oxidized at anode under anaerobic condition, the oxidation rate is very slow. The disturbance by organic matters and other factor is still unknown. Removal of ammonia inhibition in anaerobic digestion will be easily realized by this approach if anodic ammonium oxidation can be improved. Therefore, anodic ammonium oxidation is a challenging field of BES studies. In addition, the products of ammonium oxidation should be studied in more detail because nitrous oxide (N2 O) may be generated in BES under different operations. The function microbes and potential control are necessary for controlling nitrogen metabolism.

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9 Bioelectrochemical Systems for Heavy Metal Pollution Control and Resource Recovery Bo Zhang 1 , Wentao Jiao 2 , and Heming Wang 3 1 Chinese Academy of Sciences, Research Center for Eco-Environmental Sciences, CAS Key Laboratory of Environmental Biotechnology, No. 18 Shuangqing Road, Beijing 100085, China 2 Chinese Academy of Sciences, Research Center for Eco-Environmental Sciences, State Key Laboratory of Urban and Regional Ecology, No. 18 Shuangqing Road, Beijing 100085, China 3 China University of Petroleum, College of Chemical Engineering and Environment, State Key Laboratory of Heavy Oil Processing, No. 18 Fuxue Road, Beijing, 102249, China

9.1 Introduction Heavy metal pollution is a common environmental issue that occurred all over the world [1]. Conventionally, the term “heavy metal” has been used to refer to elements whose relative densities are greater than 5.0 [2]. For historical reasons, metalloids such as arsenic (As) and antimony (Sb) are also being categorized as heavy metal pollutants. Because of the adverse effects of elevated concentrations of heavy metal pollutants on the ecosystems and the health of human bodies [3], the concentrations of heavy metal pollutants in waterbodies and soils are tightly regulated. Heavy metal pollutants may enter the environment through anthropogenic activities as well as natural processes [4]. In terms of anthropogenic pathways, the mining and milling of heavy metals and the discharge of wastewater containing high concentration of heavy metal pollutants from various industries can both release heavy metal pollutants into the environment [5, 6]. On the other hand, the dissolution of rocks that are rich in heavy metals can also result in heavy metal pollution to soils and groundwater [7]. Unlike organic pollutants, a distinctive feature of heavy metal pollutants is that they cannot be degraded [8, 9]; therefore, the release of heavy metal pollutants at one specific site can have the potential to deteriorate the water quality of the entire watershed, as the pollutants will traverse along with the flow of surface water or groundwater [10, 11]. 9.1.1 Brief Review of Commonly Used Technologies for Heavy Metal Pollution Control and Their Respective Limitations To date, the most commonly used technologies for heavy metal pollution control include adsorption, flocculation-precipitation, and membrane filtration [12–14]. Bioelectrosynthesis: Principles and Technologies for Value-Added Products, First Edition. Edited by Aijie Wang, Wenzong Liu, Bo Zhang, and Weiwei Cai. © 2020 Wiley-VCH Verlag GmbH & Co. KGaA. Published 2020 by Wiley-VCH Verlag GmbH & Co. KGaA.

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In a typical adsorption-based treatment process, heavy metal laden wastewater flowed through a series of columns that were packed with adsorbent [15]. During the contact, the heavy metals were removed from wastewater and adsorbed by the adsorbent. Various materials could be used as an adsorbent, such as activated carbon [16], zeolite [17], polymers, and resins [18]. Because adsorption-based treatment processes do not require high capital investment and are easy to maintain, oftentimes, they are the preferred choice for heavy metal treatment. However, once the adsorption capacity of the adsorbent is reached, regeneration of adsorbent is required, and as a consequence, a large volume of wastewater that contains exceedingly high concentrations of heavy metals can be generated as a result [19]. The treatment of this spent regeneration solution could be quite problematic [20]. Flocculation-precipitation is another widely employed process for heavy metal wastewater treatment [13, 21]. Typically, a flocculant is added to the wastewater to induce the formation of flocculates, within which the heavy metal ions are entrapped. After the aggregation and conglomeration, the flocculates are removed in the subsequent sedimentation tank along with the heavy metal pollutants. The major advantage of this treatment technology is the same as adsorption, which is low capital investment and low maintenance requirement; however, a significant drawback of the flocculation-precipitation process is that a large volume of waste sludge will be generated, and the improper disposal of this sludge would incur secondary pollutions. In recent years, membrane-based processes are also gaining attention for the treatment of heavy metal pollutants [22–24]. Arguably, the most suitable membrane technologies for heavy metal removal are reverse osmosis and nanofiltration [25–27], which have high rejection efficiency of heavy metal ions because of the nanometer-scale pore size. Additionally, ultrafiltration-based process for heavy metal removal has also been successfully demonstrated in lab-scale study, with the addition of ligand that remarkably enlarged the size of the heavy metal ions [28–30]. However, there are two major limitations associated with membrane-based processes. Firstly, the capital and operational costs of membrane-based technologies are prohibitively high, especially considering regular maintenance and replacement of membrane modules. Secondly, the membrane-based process would inevitably generate brine solutions that contain much higher concentration of heavy metal pollutants than the feed solution, and the proper treatment of the brine could take extra effort and require additional investment.

9.1.2 Control of Heavy Metal Pollution Through (Bio)Electrochemical Processes Numerous heavy metal pollutants are redox active (see Figure 9.1), and the redox state of these heavy metals usually determines the behavior of these pollutants in the environment [31]. For example, uranium and chromium typically exist as oxyanions with a valence state of +VI in wastewater, which makes them highly soluble [32]. On the other hand, the reduced counterparts of both U(VI) and

9.1 Introduction

EH (V) pe pH 7 1.0

15

NO3– N2(g)

0.75 10 Oxic

0.5 0.25 0

5 Suboxic 0

–0.25 –5

Anoxic

Cu+ Cu0(s) SeO42– H2SeO3 Cu2+ Cu+

NO3– NO2–

Fe(OH)3(am)(s) UO2(CO3)22– Fe2+ UO2(s) CO2(g) AQDSox CH4(g) AQDH2

O2(g) H2O MnO2(s) Mn2+

H2AsO4– H3AsO3 SO42– HS– α-FeOOH(s) Fe2+

–0.5

Figure 9.1 The environmental redox potential of selected heavy metal pollutants and other inorganic electron acceptors. Source: Borch et al. 2010 [31]. Reproduced with permission of American Chemical Society.

Cr(VI) are sparingly soluble, which renders the reduction of U(VI) to U(IV) and Cr(VI) to Cr(III) promising strategies for the pollution control of uranium and chromium, respectively [33, 34]. It should be noted that the chemical reduction does not necessarily result in the immobilization of heavy metal pollutants, and the most notable example is arsenic. As(V) exists as charged oxyanions and can be easily removed through adsorption or ion-exchange process; on the other hand, As(III) typically exists as a charge-neutral molecule around neutral pH, which makes the conversion of As(III) to As(V) a prerequisite, if high removal percentage is desired [35–37]. The redox state of heavy metal could be relatively easily manipulated in electrochemical processes. Because the behavior of heavy metal pollutant is typically associated with its redox state, theoretically, the electrochemical methods should be an ideal option to treat wastewater or soils laden with heavy metal pollutants. However, to our knowledge, generally, there is a lack of interest in academia and industries in using electrochemical methods to treat heavy metal pollutions. The major reason of this lack of interest is that, compared to other treatment technologies (especially, adsorption and flocculation-sedimentation), electrochemical treatment processes are more delicate and expensive, which includes the requirement of electrodes, circuits, and specifically designed reactors. However, as more stringent discharge/water or soil quality requirements have been introduced around the globe, the interest in searching for technologies with higher effluent quality has been ever increasing. In this regard, electrochemical treatment processes are known for high effluent quality, tunable and selective removal capability, and the ability to recover and recycle the heavy metal pollutants. With these advantages, it is expected that electrochemical treatment technologies will garner more attention from scientists and practitioners in this area.

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Bioelectrochemical systems (BESs) are a suite of electrochemical devices that utilize electricigens as biocatalysts on electrode [38–40]. The electricigens grown on the surface of electrodes of BES are essentially the same as the ones that have been extensively investigated during the studies of fate and transport of heavy metal pollutants in soil and groundwater; therefore, in this regard, BES is very suitable to be employed in the treatment of heavy metal pollutions [41–44]. Additionally, BES also possess the advantages of traditional electrochemical process, among which the most desired one is tunable electrode potential [45, 46], which would allow practitioners an extra operational parameter to selectively and/or sequentially remove heavy metal pollutants. Such flexibility is highly desirable in the application, as it would pave the way for the recovery and reuse of heavy metals. In this chapter, we will firstly revisit the principles of BES in the context of pollution control; we will then introduce the research progresses that have been made in the field of heavy metal pollution control using BES; thirdly, we will share our outlook on how BES could be utilized as a platform technology for enhanced heavy metal pollution control.

9.2 BES and its Application in Heavy Metal Pollution Control 9.2.1

Configuration of BES

Typically, BES refers to a suite of electrochemical devices that rely on electrodeactive microorganisms (also referred to as electricigens or exoelectrogens) to catalyze anode or cathode reactions [47, 48]. In the past two decades, BES has received extensive research attention because of its versatility and its prospects in recovering energy and resources from wastewater and soil. To date, the concept of BES encompasses various bioelectrochemical devices. However, microbial fuel cells (MFCs), as one of the earliest configuration of BES, arguably remain to be the most extensively studied ones (see Figures 9.2 and 9.3) [39, 40]. In a typical MFC, electricigens are cultivated on anode, which is used as the terminal electron acceptors while metabolizing substrates (i.e. organic matters in wastewater). The electrons collected by anode are transferred through external circuit to cathode and used to reduce electron acceptors (such as oxygen or heavy metal ions). Typically, the working anode potential is higher than −0.3 V (vs. standard hydrogen electrode) [48], which limits the cathodic reactions to those with redox potential higher than this value. This limitation could be remedied by adding an external bias to the MFC, turning the MFC to a microbial electrolysis cell (MEC, see Figure 9.2) [49–51]. With the aid of the supplemented bias, the cathode potential could be driven to the more negative direction; hence, some electron acceptors that were unable to be utilized in MFC can now be used as electron acceptor on the cathode of MEC. For example, H2 O and CO2 can be used as electron acceptor at the cathode of MEC but not by that of MFC, and the reduction products of H2 O and CO2 are hydrogen and methane, respectively

9.2 BES and its Application in Heavy Metal Pollution Control

V

MFC mode

(b)

V

MFC mode

Resister

Resister MEC mode

MEC mode

Resister

Power source Resister e

CO2 H+ Anode

IEM

e–



Potentiostat e–



e

CO2

MEox

H+

H+

MEred Abiotic cathode

Organics

Anodic bacteria

IEM

(a)

MEox +

H

MEred Biotic cathode

Organics Anode

Cathodic bacteria

Figure 9.2 The schematic diagram of different BESs in controlling heavy metal laden wastewater with abiotic cathode (a) or biotic cathode (b). In MFC mode, no external power supply is needed; in MEC mode, power supply is supplemented to the external circuit. IEM, ion exchange membrane.

V

(a)

V

(b)

Resister

Resister

e–

e–

e–

e–

Anode

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MEox

H+ MEred Air cathode

CO2 H+

O2 Anode

Organics

IEM

IEM

H2O CO2 H+

MEox

H+

MEred Non air cathode

Anodic bacteria

Figure 9.3 The schematic diagram of different BESs in controlling heavy metal-contaminated soils with air-cathode (a) or non-air-cathode (b). IEM, ion exchange membrane.

[52, 53]. Therefore, it is obvious that the scope of the BES is greatly extended by incorporating MEC, and the electron acceptors with lower redox potentials can be reduced by the cathode of MEC. As we will discuss in detail in the following part of this chapter, the electrochemical reduction of a host of heavy metal pollutants has the effect of greatly reducing the solubilities of heavy metals in water; therefore, electrochemical reduction by BES, in these circumstances, is a powerful strategy in treating heavy metal polluted wastewater or soils. From this perspective, MEC has wider applicability in treating heavy metal pollution, as heavy metal pollutants with low redox potentials could be removed by manipulating cathode potential. BES can also be categorized based on the type of catalysts used for cathode reactions. On the one hand, noble metal catalysts (such as Pt and Pd), nanomaterials, and modified bulk carbon materials have all been tested as catalysts in BES

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[40, 54–56]. When these materials were used as cathode catalyst, the respective BES was considered to have an abiotic cathode. On the other hand, BES could also rely on electrochemical-active biofilms to catalyze the cathode reactions, and in these circumstances, the BES is considered to have a biocathode [45, 57]. Typically, higher current could be obtained from the abiotic cathode, but the catalysts may gradually experience a decrease in catalytic activity because of fouling or even poisoning. However, biocathode would generally remain active during the operation as the biofilms regenerate in metabolism, which makes the process long-lasting and sustainable. Separator is another important component of BES. In the early studies, a single piece of semipermeable ion exchange membrane (IEM) was typically used to separate the anode and cathode chamber [48, 58]. In a later study, Cao et al. developed the so-called microbial desalination cell (MDC) by adding an additional complementary piece of IEM between the anode and cathode chambers [59]. This modification essentially formed a middle chamber in MFC, allowing the salt solution in the middle chamber to be desalinated, in a manner similar to an electrodialysis cell. Attempts have also been made to replace the IEM with other types of membranes. Zhang et al. firstly developed the so-called osmotic microbial fuel cell (OsMFC) by replacing IEM with forward osmosis membrane [60]. With this modification, the salinity gradient between wastewater in anode chamber and the salty water in cathode chamber can be utilized, and a spontaneous forward osmosis process would drive H2 O molecules from the anode chamber to the cathode chamber, which has the net effect of water recovery. Further attempts were made to incorporate the forward osmosis process into MDC, and the developed osmotic microbial desalination cell (OsMDC) achieved the goal of simultaneous water recovery and partial salt removal [61]. Because elevated concentrations of heavy metal pollutants generally have toxic effect on microorganisms, to date, most studies on heavy metal pollution control using BES technologies commonly employ IEMs to separate the anode chamber and cathode chamber, in order to limit the possible negative effect of heavy metals on anodic microorganisms. 9.2.2

BES Application in Treating Heavy Metal Laden Wastewater

In the following sections, we will review the application of BES in heavy metal pollution control based on the configurations of BES used. The vast majority of studies in this area focused on using BES to convert the heavy metal ions from oxidized state to reduced state (i.e. reduction reaction). In line with this trend, we will review the application of BES in two categories, namely, heavy metal reduction using abiotic cathode and heavy metal reduction using biotic cathode (biocathode). 9.2.2.1

BES with Abiotic Cathode

As mentioned earlier, the equilibrium potential of anode is typically near −0.3 V (vs. standard hydrogen electrode); therefore, for the reduction reaction to take place in MFC, the redox potential of heavy metal should be at least higher (more positive) than this value. With this limitation, only a selected group of heavy metal ions can be removed in MFC mode reactors.

9.2 BES and its Application in Heavy Metal Pollution Control

To date, the heavy metal pollutants can be directly reduced include Au(III), Ag(I), V(V), Cr(VI), and Se(IV) (see Table 9.1). Choi et al. demonstrated that over 99% of Au(III) could be removed and recovered in a two-chamber MFC [62]; in a separate study, they further demonstrated that Ag(I) could be reduced onto the abiotic cathode in MFC with a high removal efficiency [69]. Using Au(III) and Ag(I) as cathode electron acceptor, the power density of two-chamber MFCs was 6.58 and 4.25 W/m2 , respectively. As proof-of-concept studies, the initial concentrations of Au(III) and Ag(I) were c. 10 mM at the beginning of the reaction. This concentration was relatively high compared to real wastewater, and the hydraulic retention times (HRTs) were relatively long, which were 25 and 8 hours, respectively, to achieve the complete removal of Au(III) and Ag(I). However, for the MFC to be applied in application, the concentration of electron acceptor is likely much lower than the value used in these studies, and the retention time should also be shortened to reduce the capital cost of the treatment process. Therefore, the power generation from the MFC would likely be significantly reduced. In addition to Au and Ag, pentavalent vanadium and hexavalent chromium could also be removed from solution by abiotic cathode in MFC. The formal redox potentials of V(V) and Cr(VI) were 0.99 and 1.33 V, respectively, which make them ideal candidates to be used as terminal electron acceptor on cathode. However, the removal efficiencies for these two types of pollutants were lower compared to Au and Ag. With an initial concentration of c. 9.8 mM, only 25% V(V) was removed after 72 hours [63]; similarly, the removal efficiency for Cr(VI) was only 67.9% in a later study by the same group [64]. While the experimental conditions (such as reactor configuration, HRT, external resistance, etc.) were different in each study, the gap in removal efficiencies between the group of Au and Ag and the group of V and Cr clearly indicated that there is no one solution for all type of process for heavy metal reduction in MFC, the specific chemical behavior of the target heavy metal pollutants should be taken into consideration, so that the proper cathode material and reactor configuration could be selected to achieve the desirable degree of removal efficiency. The above-mentioned studies used two-chamber MFCs to treat heavy metal laden wastewater. Traditionally, there is an inherent drawback of two-chamber MFCs; that is, they often suffer from high internal resistance, which in turn prohibits the flow of electrons. Therefore, researchers begin to consider using single chamber to treat the heavy metal wastewater. Catal et al. demonstrated that Se(IV) was reduced to elemental Se in a single-chamber MFC with a recovery efficiency of 99% Se(IV) [78]. However, because Se is toxic toward microorganisms, the power generation from the MFC decreased from 2.9 to 2.2 W/m2 as Se(IV) concentration increased from 0.31 to 0.95 mM. This case demonstrated that when using BES to treat heavy metal laden wastewater, the toxicity of the heavy metals should be considered, and we consider that a separator is indispensable when BES was employed to treat toxic metals. In addition to the toxicity issue, Se(IV) also behaved like a competing electron acceptor for the anode, which would inevitably result in lower current and power generation from MFC. Taken these two issues together, we deem it is necessary in certain scenarios to incorporate a separator into the BES for efficient heavy metal removal.

239

Table 9.1 Heavy metal pollution control using abiotic cathode.

No.

Operation mode

Supplemented bias/cathode potential (NHE)

Time (h)

Reduction efficiency

Power/ energy

References

1.002

MFC

0

25

99.89 ± 0.00% (for 200 mg/l Au(III))

890–6580 mW/m2

Choi et al. [62]

Standard redox potential (vs. NHE, V)

Metal

Reaction

1.

Au(III)

AuCl−4

2.

V(V)

VO+2 + 2H+ + e− = VO2+ + H2 O

0.991

MFC

0

72

25.3 ± 1.1%

572.4 ± 18.2 mW/m2

Zhang et al. [63]

3.

V(V)

VO+2 + 2H+ + e− = VO2+ + H2 O

0.991

MFC

0

240

67.9 ± 3.1%

970.2 ± 20.6 mW/m2

Zhang et al. [64]

Cr(VI)

+ − Cr2 O2− 7 + 14H + 6e = 2Cr3+ + 7H2 O

1.33

4.

Cr(VI)

+ − Cr2 O2− 7 + 8H + 6e = Cr2 O3 + 4H2 O Cr2 O3 + 6H+ = 2Cr3+ + 3H2 O

1.33

MFC

0

10 or 23.3

Cr6+ (99.5%) total Cr (66.2%)

1600 mW/m2 (204 mg/l Cr6+ )

Li et al. [65]

5.

Cr(VI)

+ − Cr2 O2− 7 + 14H + 6e = 2Cr3+ + 7H2 O

1.33

MFC

0

150

74.6–100%

150 mW/m2

Wang et al. [66]

6.

Cr(VI)

+ − Cr2 O2− 7 + 14H + 6e → 2Cr3+ + 7H2 O 2Cr3+ + 7H2 O → 2Cr(OH)3 (s) + 6H+ + H2 O (6.5 < pH < 10)

1.33

MFC

0

4, 5

12.4–20.6 mg/g VSS/h

6.8–15 W/m3

Huang et al. [67]

7.

Cr(VI)

Cr6+ → Cr3+

1.33

MFC

0

7

2.0 ± 0.1–2.4 ± 0.2 mg/g VSS/h

0.3–2.4 W/m3

Huang et al. [68]

8.

Ag(I)

Ag+ + e− = Ag

0.799

MFC

0

8

99.91 ± 0.00– 98.26 ± 0.01%

4250 mW/m2

Choi et al. [69]



+ 3e = Au + 4Cl



75.4 ± 1.9%

9.

Ag(I)

Ag+ + e− = Ag

0.799

[AgS2 O3 ]− + e− = Ag + S2 O3 2−

0.250

MFC

0

2 or 36

95%

109 mW/m2

Tao et al. [70]

35 mW/m2

10.

Cu(II)

Cu2+ + 2e− = Cu

0.286

MFC

0

7

99.88% (anaerobic), 99.95% (aerobic)

430 mW/m2 (anaerobic), 800 mW/m2 (aerobic)

Heijne et al. [71]

11.

Cu(II)

Cu2+ + 2e− = Cu

0.337

MFC

0

11

60.1–99.9%

96.2–319 mW/m2

Wang et al. [72]

12.

Cu(II)

Cu2+ + 2e− = Cu 2Cu2+ + H2 O + 2e− = Cu2 O + 2H+ Cu2+ + Cl− + e− = CuCl

NA

MFC

0

480, 672

92%, 48%

NA

Tao et al. [73]

13.

Cu(II)

Cu2+ + 2e− = Cu 2Cu2+ + H2 O + 2e− = Cu2 O + 2H+ Cu2 O + 2e− + 2H+ = 2Cu + H2 O 4Cu2+ + SO4 2− + 6H2 O = CuSO4 ⋅ 3Cu(OH)2 + 6H+

0.337, 0.207, 0.059

MFC

0

20, 360

>96%

339 mW/m3

Tao et al. [74]

14.

Cu(II)

Cu2+ → Cu2 O Cu2+ → Cu2 O + Cu CuSO4 → Cu4 (OH)6 SO4

NA

MFC

0

144

70%

314 mW/m3

Tao et al. [75]

15.

Fe(III)

Fe3+ + e− = Fe2+ Fe2+ → oxy(hydroxi)des

0.77

MFC

0

168

10–>99%

8.6 ± 2.3 W/m3

Lefebvre et al. [76] (Continued)

Table 9.1 (Continued)

No.

Metal

Standard redox potential (vs. NHE, V)

Reaction 2+



2+

Operation mode

Supplemented bias/cathode potential (NHE)

Time (h)

Reduction efficiency

Power/ energy

References 2

16.

Hg(II)

2Hg + 2e = Hg2 Hg2 2+ + 2e− = 2Hg Hg2+ + 2e− = Hg Hg2 2+ + 2Cl− = Hg2 Cl2 Hg2 Cl2 + 2e− = 2Hg + 2Cl−

0.911, 0.796, 0.851, 0.268

MFC

0

5–10

89.5–99.3%

433.1 mW/m

Wang et al. [77]

17.

Se(IV)

− SeO2− 3 + 4e → Se(s)

NA

MFC

0

42 or 72

99%

2900 mW/m2 (acetate) 1500 mW/m2 (glucose)

Catal et al. [78]

18.

Cd(II)

Cd2+ + 2e− = Cd

−0.4

MFC

0

50

90%

3600 mW/m2

Abourached et al. [79]

Zn(II)

Zn2+ + 2e− = Zn

−0.764

19.

Zn(II)

Zn2+ permeation

−0.764

MFC

0

72

94 ± 4%

0.233 mW

Fradler et al. [80]

MEC

1.0 V

90

99.2 ± 0.1%

NA

Luo et al. [81]

>3.89 W/m2

Choi et al. [82]

20.

Cu(II)

Cu2+ + 2e− = Cu(0)

0.34

Ni(II)

Ni2+ + 2e− = Ni(0)

−0.25

97%

97.0 ± 1.3%

Fe(II)

Fe2+ + 2e− = Fe(0)

−0.44

21.

Cr(VI)

Cr6+ + 3e− = Cr3+

1.33

Cd(II)

Cd2+ + 2e− = Cd(s)

−0.403

MEC

>0.128 V

22.

LiCoO2

Co3+ + e− = Co2+

1.61

MFC

0

Co(II)

Co2+ + 2e− = Co(0)

−0.232

MEC

Up to 0.4 V

Source: Adapted from Wang and Ren 2014 [84].

97.0 ± 1.8% MFC

0

60

>7.51 ± 0.06% >89.73 ± 0.28%

NA

6

2% (% leached from solid)

5.0 W/m3 (pH = 1.0)

40%

NA

Huang et al. [83]

9.2 BES and its Application in Heavy Metal Pollution Control

There is one scenario in which separatorless configuration would be inducive to the heavy metal removal in BES. In wastewater that contains Cd and Zn together with sulfate, it has been shown that Cd and Zn could be deposited to the surface of electrode (in this case, the electrode is anode) through biosorption and sulfide precipitation, with a modest power output of 3.6 W/m2 [79]. The feasibility of this process partly lies on the fact that Zn and Cd have limited toxicity to the electrode biofilm, as evident by the maximum tolerable concentrations (MTCs) of Cd and Zn were 200 and 400 mM, respectively. However, because sulfate commonly exists in heavy metal wastewater, especially wastewater from mining industries, this scenario should also be considered and exploited during the selection of reactor configuration. For pollutants whose reduction potential were relatively lower, an external bias is usually supplemented to facilitate their reductive removal. Luo et al. studied the removal of Cu(II), Ni(II), and Fe(II) from a synthetic acid mining drainage effluent [81]. The reduction potential of Cu(II), Ni(II), and Fe(II) were +0.34, −0.25, and −0.44 V, respectively. While the reduction of Cu(II) could be driven directly by MFC, the reduction of Ni(II) and Fe(II) cannot be driven by electrons derived from bioanode because their reduction potentials were too close to the equilibrium potential of the anode. On the other hand, with the externally added bias of 1 V, Luo et al. successfully achieved high rate removal of Cu(II) and Ni(II), whereas Fe(II) was removed at a lower rate. The reported removal rates were 560, 204, and 168 mg/l/d for Cu(II), Ni(II), and Fe(II), respectively. Another interesting point from this study was that hydrogen gas was produced simultaneously in the cathode chamber because of the relatively large applied potential and the coating of Pt on the surface of the cathode. It was inferred that H2 migrated through the bipolar membrane used in this study because the calculated coulombic efficiencies were larger than 100%. If the only purpose of the BES is to treat heavy metal pollutants, then a bias much lower than 1 V should suffice. In fact, a natural choice in this circumstance is to use MFC as the power supply to drive the MEC for heavy metal recovery. Choi et al. demonstrated the possibility of using a Cr(VI)-reducing MFC as power supply to drive the Cd(II)-reducing MEC. In their experimental setting, the Cd(II) removal efficiencies could be as high as 94.43% ± 0.17%, 93.30% ± 0.74%, and 89.73% ± 0.28% for initial Cd(II) concentrations of 50, 100, and 200 ppm, respectively [82]. The Cr(VI)-reducing MFC provided at least 0.128 V external bias to the Cd(II)-reducing MEC; however, to maintain a stable power output, the MFC was operated at elevated Cr(VI) concentrations in the cathode side, which led to low removal efficiencies for Cr(VI). The Cr(VI) removal efficiencies after 60 hours of operation was 13.95% ± 0.73%, 10.39% ± 0.64%, and 7.51% ± 0.06% for initial Cr(VI) concentration of 200, 400, and 800 ppm, respectively. Huang et al. also investigated the cobalt leaching and recovery in MFC–MEC process [83]. The LiCoO2 was firstly used as the electron acceptor in the cathode of MFC. Owing to its relatively positive reduction potential, Co(III) in LiCoO2 was firstly reduced in the cathode of MFC to Co(II). The Co(II) containing solution was then transferred to cathode of MEC for further conversion of Co(II) to Co(0). Because the reduction potential of Co(II) was very close to the anode potential, the cobalt-leaching MFC was employed as power source

243

244

9 Bioelectrochemical Systems for Heavy Metal Pollution Control and Resource Recovery

to drive the Co(II) reduction in MEC. After process optimization, a leaching rate of 46 ± 2 g/Co⋅l/h and a reduction rate of 7.0 ± 0 g/Co⋅l/h were achieved. It is worth noting that the Co(III)-leaching MFC and Co(II)-reducing MEC had different optimal operating conditions in terms of catholyte composition. Although high solution conductivity is preferred in the MFC, the MEC should be better operated at a lower conductivity. In fact, the increase of catholyte in MEC would significantly reduce its internal resistance and led to high current density, and concomitantly, the operating voltage of MFC would decrease because of the increased current in the circuit, which in turn would lower the reduction rate in the MEC. 9.2.2.2

BES with Biocathode

In addition to abiotic cathode, the reductive removal of heavy metal pollutants can also be achieved by using biocathode (see Table 9.2). In fact, biocathode has been employed in BES for the treatment of various types of pollutants, such as nitrate, nitroaromatic compounds, and antibiotic [57, 88, 89]. Compared to abiotic cathode, the most distinctive advantage of biocathode is that the electrode-active biofilm developed on the surface of cathode harbors a host of enzymes, which could catalyze the reductive transformation of different types of pollutants in the wastewater. Gregory and Lovley firstly attempted to use biocathode for the reductive immobilization of U(VI) [85]. In their study, a graphite electrode poised at −0.5 V (vs. Ag/AgCl) was found to be able to rapidly remove the U(VI) in solution; however, the removed U(VI) returned to the solution once the bias was removed from the graphite electrode. This fact indicated that the removal of U(VI) by bare graphite electrode should be ascribed to electroadsorption, instead of electrochemical reduction. On the other hand, when the dissimilatory metal reducing bacterium Geobacter sulfurreducens was added to the reactor, relatively slow yet steady U(VI) concentration decrease was observed with the same applied potential; more importantly, after the graphite electrode was stopped being poised, U(VI) did not return to the solution phase, which was a clear sign that U(VI) was reduced to U(IV). Numerous studies have demonstrated that dissimilatory metal reducing bacteria (such as Geobacter spp. and Shewanella spp.) could enzymatically reduce U(VI) to U(IV), using the redox-active outer membrane c-type cytochromes as the biocatalysts [90–92]. The study of Gregory and Lovley illustrated that the same enzymatic pathway could be exploited by biocathode to facilitate the reduction of U(VI) to U(IV), which has the effect of immobilizing the uranium pollution. Chromium is another heavy metal pollutant whose removal was extensively investigated using biocathode BES. In a study published in 2010, Huang et al. demonstrated that Cr(VI) could be removed using a biocathode MFC. The reported chromium removal efficiency was 2.4 ± 0.2 mg/Cr g/VSS h; however, this removal efficiency was not higher than that of conventional aerobic or anaerobic Cr(VI) removal processes [68]. The advantage of Cr(VI) removal using MFC, on the other hand, is power production. By using Cr(VI) as an electron acceptor in the cathode, a power density of 6.9 A/m3 was achieved.

Table 9.2 Heavy metal pollution using biocathode BES.

No. Metal

Reaction −

Supplemented Standard bias/cathode redox potential Operation potential (NHE) (vs. NHE, V) mode

Time Reduction (h) efficiency

Power/ energy

NAa)

MEC

−0.5 to −0.7 V (set potential)b)

24

NA

NA

Gregory et al. [85]

MFC

NA

7

99.3%

Max power density: 2.4 ± 0.1 W/m3

Huang et al. [68]

5

85 ± 1% 78 ± 1%

NA

Hou et al. [86]

1.

U(VI)

U(VI) + 2e = U(IV)

2.

Cr(VI)

Cr(VI) + 3e− = Cr(III) 1.33

3.

Cd(II) Cr(VI) Cd(II) + 2e− = Cd(0) −0.40 1.33 Cr(VI) + 3e− = Cr(III)

MEC

Applied voltage: 0.5 V; Cathode potential: c. −0.6 V

4.

Cu(II)

Cu(II) + 2e− = Cu(0)

0.34

MFC

Cathode 4 potential: −0.110 ± 0.002 V

Complete Energy (1.24 ± 0.01 mg/l h)c) production: 0.6 kWh/kg-Cr

Cr(VI)

Cr(VI) + 3e− = Cr(III) 1.33

MFC

Cathode potential: −0.074 ± 0.001 V

Complete (1.07 ± 0.01 mg/l-h)

Energy production: 0.3 kWh/kg-Cu

Cd(II)

Cd(II) + 2e− = Cd(0)

MEC

Cathode potential: −0.610 ± 0.004; Applied voltage: 0.5 V

Complete (0.98 ± 0.01 mg/l-h)

Energy consumption: 10.5 kWh/kg-Cd

−0.40

a) U(VI) may form various complexes with inorganic ligands such as bicarbonate, and different complexes may have different reduction potentials. b) The cathode was poised at −0.5 to −0.7 V (vs. Ag/AgCl, saturated KCl) during the experiment. c) The removal rate of heavy metal pollutants was reported in the parenthesis.

References

Huang et al. [87]

246

9 Bioelectrochemical Systems for Heavy Metal Pollution Control and Resource Recovery

This power generation should largely be ascribed to the high redox potential of Cr(VI)/Cr(III) pair, and if both power generation and Cr(VI) removal were desired, then biocathode MFC could be a viable treatment method. Although it is important to study the removal of single pollutant in the waste stream in the proof-of-concept stage, it is not uncommon that multiple heavy metal pollutants presented simultaneously in the real wastewater. In such cases, in order to better understand the removal process, it is important to understand (i) how the coexistent heavy metals affect the removal of individual pollutant and (ii) the characteristics of microbial community of biocathode during multiple heavy metal reduction. Hou et al. studied the simultaneous removal of Cd(II) and Cr(VI) in a biocathode MEC [86]. Compared to the controls in which Cd(II) and Cr(VI) presented individually, it was found that biocathode exhibited a higher Cd(II) removal rate but a lower Cr(VI) removal rate. Using fluorescent probes, this study revealed that Cd and Cr were partitioned in different parts of the electrode-active bacteria of biocathode. Although more Cd was associated with the cell membrane, it was found that Cr(III), which was the reduction product of Cr(VI), was more concentrated within the cytoplasmic spaces. Such difference in the active site of heavy metal removal could be a reason as to why the Cd(II) removal was enhanced while the Cr(VI) removal was impeded, when these two heavy metal presented simultaneously. 9.2.3

BES Application in Controlling Heavy Metal Polluted Soils

In addition to being studied as a promising technology for water pollution control, BES has also been investigated as a viable option for soil pollution remediation (Figure 9.3). Habibul et al. explored the possibility of using an air-cathode MFC for Cd and Pb removal from polluted soil [93]. In their experimental setting, the cathode chamber of an MFC was packed with polluted soil; upon current generation, the electric field induced by anode and cathode reactions drove the positively charged Cd(II) and Pb(II) ions migrate toward the cathode of the reactor. At the end of the experiment, 31.0% and 44.1% of Cd and Pb were removed, respectively. The remediation experiment lasted for 143 and 108 days for Cd and Pb, respectively. The current densities (normalized to the anode surface area) generated by the MFC were only 0.6–1.2 mA/cm2 for Pb(II) and 0.7–1.5 mA/cm2 for Cd(II). In essence, the remediation technology proposed by Habibul et al. is very similar to the electrokinetic remediation; although the current densities reported in their study were lower than that of typical electrokinetic remediation process, no extra power was needed in MFC-based remediation technology, which inevitably led to a more sustainable system. Other researchers also made attempt to incorporate the electrokinetic remediation with other types of BES. Chen et al. constructed a three-chamber BES and packed the middle chamber with Zn and Cd polluted soil [94]. This experimental device was similar to the MDC in a sense that the ions in the middle chamber were removed to maintain the charge balance in the anode chamber and cathode chamber. After 78 days of operation, 25% and 18% of the Zn(II) and Cd(II)

9.3 Outlook and Concluding Remarks

were removed from the polluted soil in middle chamber. Compared to the study by Habibul et al., the current density of the three-chamber soil MFC was lower, which was c. 0.04 mA/cm2 . We believe that this partly could be explained by the higher internal resistance of the three-chamber MFC. While the cube-shaped reactor can borrow the existing designing and operation experiences from the previous BES studies, the implementation of such devices in situ would inevitably require more excavation efforts. In order to make soil MFC more engineering-wise viable for in situ remediation purpose, Wang et al. designed a tube-shaped MFC and studied its efficacy in removing Cu(II) from the polluted soil [95]. The configuration of this tube soil MFC was similar to the benthic MFC, with the anode placed at the lower part of the column and cathode placed at the top layer of soil column. A distinctive advantage of such tube-shaped configuration is that devices and experiences from hole-boring practice could be directly applied, which would greatly facilitate the implementation of this device in field. Upon current generation, the Cu(II) ions would migrate from the bottom of the soil column to the top layer because of the charge imbalance caused by anode and cathode reactions, and be concentrated near the cathode. Partly because of the high internal resistance caused by the soil column, the current generated in this tube soil MFC was low, with the highest current density of 0.051 mA/cm2 (normalized to the surface area of the separator). After a duration of 56 days, the total Cu content on the cathode surface increased from 8.10 to 278.25 mg/kg; additionally, clear concentration gradient of soluble Cu was observed, with the segments near to the cathode had higher soluble Cu concentration. These data strongly corroborated the hypothesis that copper migrates to the cathode under current generation. However, further analysis revealed that not all the copper was deposited on the cathode surface, rather than that, larger portion of the Cu(II) were still within the soil matrix. If complete removal of Cu is the goal, then further elongation of remediation period is required. In addition, it would be interesting to investigate the microbial community shift along the soil column because a dynamic change of the Cu concentration within different segments of the soil column existed, and it is expected that the microbial community would shift accordingly to cope with the stress induced by toxic Cu. Such information would not only help us to understand the Cu(II) removal in soil MFC but would contribute to our knowledge regarding the fate and transport of Cu in polluted sites.

9.3 Outlook and Concluding Remarks While using BES for heavy metal removal has been extensively studied in the past decade, this technology is still at the early stage, and significant improvements are required to enable BES as a viable and attractive option for practitioners when designing full-scale treatment processes. First of all, it appears that operating BES in MEC mode is more suitable for the purpose of pollution control. Although Cu(II) and Cr(VI) have been demonstrated to be used as electron acceptor in the cathode reaction of MFC, the

247

248

9 Bioelectrochemical Systems for Heavy Metal Pollution Control and Resource Recovery

produced power densities were generally low. On the other hand, pollutants with low redox potential, such as Cd(II), can only be removed in the cathode chamber of MEC. Considering that waste stream typically contained more than one type of pollutants, and that lowering the cathode potential is thermodynamically more favorable for higher percentage of removal, it is therefore suggested that future studies should place the focus on developing MEC processes for heavy metal pollution control. Secondly, the niche of the BES application for heavy metal pollution control needs to be carefully selected. Although it has been demonstrated that BES could remove various redox-active heavy metals, the removal rate is generally not higher than conventional processes. Considering that certain waste streams may contain exceedingly high concentration of toxic heavy metals or have extreme pH values, it is suggested that the BES-based treatment process should be utilized in coordination with other treatment processes. Ideally, the influent stream should contain moderate level of heavy metal pollutants with near-neutral pH; in such scenario, the advantage of enzymatic catalytic reduction provided by biocathode could be utilized for efficient reductive removal of heavy metals. Thirdly, similar to other BES studies, the investigation of heavy metal removal in BES has been primarily conducted with lab-scale reactors. The enlargement studies are required to better understand the effect of reactor configuration, electrode material on the treatment efficiencies. For applications that involve biocathode, there is a scientific question of what the microbial community would be in treating real heavy metal wastewater, as well as the engineering question of how to maintain the activity of electrode-active bacteria during temperature and waste concentration fluctuations.

Acknowledgments This study was financially supported by Beijing Natural Science Foundation (No. 8184085), Natural Science Foundation of China (Nos. 21806185 and 21806176), and Science Foundation of China University of Petroleum-Beijing (No. 2462015YJRC016).

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10 External Electron Transfer and Electrode Material Promotion Fanghua Liu, Hengduo Xu, and Jiajia Li Chinese Academy of Sciences, Yantai Institute of Coastal Zone Research, Key Laboratory of Coastal Biology and Utilization, CAS Key Laboratory of Coastal Environmental Processes and Ecological Remediation, 17 Chunhui Road, Laishan District, Yantai, Shandong 264003, China

The phenomenon of electron exchanged between microorganisms and electrodes has been widely studied and leads to a promising and novel technology: microbial electrochemical. As many microorganisms possess bidirectional electron transfer capability, i.e. electrons transfer from microorganisms to electrodes that are developed for the microbial fuel cell (MFC) and electrons transfer from electrodes to microorganisms that are [1] developed for the microbial electrosynthesis (MES) or bioelectrosynthesis of useful products from CO2 . Actually, the biological capacity to uptake and transfer electrons among solid donors is a widespread process that promotes biogeochemical cycles in nature. A significant breakthrough in the extracellular electron transfer (EET) was the finding that nonfermentable substrates such as acetate could be oxidized to carbon dioxide accompanied by electron transfer to an electrode acting as the electron acceptor [2, 3]. Opposite to the electron transfer to electrode, the history of electron transfer from electrode to microorganism is relatively short. The direct microbial electron flow to electrode was first reported by Gregory et al. [4], where they demonstrated that graphite electrode could act as electron donors for microbial respiration. Meanwhile, electricity has been severed as power to stimulate microbial metabolism for several decades. The products generated from bioelectrosynthesis process seriously depend on electricity through EET (i.e. external electron transfer) from electrodes to microorganisms because electron delivered from the electrode is the only energy source for microorganisms reducing CO2 into valued-added products [5]. Therefore, EET significantly influences the bioelectrosynthesis process. In order to promote commercialization of bioelectrosynthesis, one key feature is promoting electron exchange at the electrode surface while maintaining low costs. There subsequently occurred numerous studies on a wide diversity of electrode materials that can influence the electron transfer efficiency or fermentation patterns of microorganisms [6, 7] (Figure 10.1). For example, Zhang et al. demonstrated that carbon cloth electrodes modified with positively charged molecules, chitosan, or cyanuric chloride could highly increase acetate Bioelectrosynthesis: Principles and Technologies for Value-Added Products, First Edition. Edited by Aijie Wang, Wenzong Liu, Bo Zhang, and Weiwei Cai. © 2020 Wiley-VCH Verlag GmbH & Co. KGaA. Published 2020 by Wiley-VCH Verlag GmbH & Co. KGaA.

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e–

e– e–

Nanowire

e– c-type cytochrome

Direct electron transfer

e– Medred

Medox

Endogenous mediator shuttled Artificial mediator shuttled

e–

Indirect electron transfer

H+ H2 Anode

Interspecies H2 transfer

Cathode

Figure 10.1 Electron transfer mode on the cathode in bioelectrosynthesis.

production with six- to sevenfolds compared to unmodified electrode [8]. Therefore, it is obvious to predict that an appropriate electrode material will enhance the bioelectrosynthesis performance. The primary focus of this chapter is to discuss the EET strategies employed by autotrophic microorganisms to uptake electrons from electrodes in bioelectrosynthesis systems. We also review recent process achieved toward the development of electrodes that improves the bioelectrosynthesis performance.

10.1 External Electron Transfer The microorganisms that are able to transfer electrons extracellularly often have a cytoplasmic membrane on the cell envelopes. This cytoplasmic membrane acts as a primary barrier to the external environment and a center for electron transfer, which plays an important role in microbial electricity production. However, some other structural components such as peptidoglycan and the outer membrane exist on the cell envelope of microorganisms. These components are electrically nonconductive and physically impermeable to electron acceptors [9]. The electron transfer between redox carriers in the cytoplasmic membrane and extracellular electron acceptors is often called as microbial EET. EET involves direct electron transfer (DET) and indirect electron transfer (IDET) and has attracted wide interest in relation to bioelectrosynthesis. In particular, external supply of electrons through electricity drives the process of reducing CO2 and further forms multicarbon products such as methane and other biofuels. Therefore, exploring the EET mechanisms involved in bioelectrosynthesis may result in important breakthroughs in the effort to promote the electron transfer rate from the cathodes to the microbial catalyst. We will discuss the two modes of EET in the following section.

10.1 External Electron Transfer

10.1.1

Direct Electron Transfer

DET is ideal in EET in which microorganisms uptake electrons directly from the surface of the cathode without electron mediators such as H2 and formate. The DET process often needs a cathode potential that is relatively more negative compared to IDET, which results in a high current density [10]. The microorganisms should physically contact with cathodes during DET process and form biofilms on the cathodes, which take advantages of directly catalyzing CO2 reduction. In addition, the biofilms on the cathode remains a long time. However, the diffusion of substrate and product will be limited among the biofilms. In particular, DET process may be restricted by the surface area of cathodes in bioelectrosynthesis systems at industry scale. The DET process is thought to transfer electrons between cytoplasmic membranes of microorganisms and cathodes by the aid of a network of redox and structural proteins. Some of these proteins have been deeply studied in a few model microorganisms. These proteins usually establish pathways that electrically and physically connect intracellular electron transfer with redox transformations of extracellular electron acceptors [11]. It should be noted that the components of these pathways are phylogenetically diverse and thus cannot always be identified through conventional analyst technology such as genomic tool. Therefore, the development of a mechanistic understanding of microbial EET pathways requires the identification and functional characterization of their components. However, there are few literature studies reported on the redox proteins located on the surface of cells, and the corresponding electron transfer mechanism is still unknown. The two relatively extensively studied microorganisms for EET mechanisms are Geobacter and Shewanella, which also names as metal-reducing and divided as gram-negative bacteria. These model microorganisms have c-type cytochrome (c-Cyt) that is of paramount importance in EET process. Shewanella oneidensis MR-1 is the first identified microorganisms capable of transferring electrons to Fe3+ or Mn4+ as terminal electron acceptors [12]. This bacterium consists of six multihaem c-Cyts, i.e. CymA, Fcc3, MtrA, MtrC, OmcA, and small tetrahaem cytochrome (STC) and the porin-like outer membrane protein MtrB. All these c-Cyt directly take part in DET process. Protein purification and characterization has revealed that quinol in the cytoplasmic membrane is firstly oxidized by CymA, which further transfers the released electrons to the periplasmic Fcc3 and STC [13, 14]. After that, Fcc3 and STC transfer electron to MtrA that in the outer membrane of cells. MtrA, MtrB, and MtrC form a trans-outer membrane protein complex that transfers electrons from the periplasmic proteins to the bacterial surface. At last, MtrC and OmcA physically interact with each other on the bacterial surface and transfer electrons directly to electron acceptors such as minerals and electrodes through solvent-exposed haems. For Geobacter sulfurreducens, multihaem c-Cyts also directly participate in electron transfer from the cell envelope to electron acceptors. The known multihaem c-Cyts mainly consist of the putative quinol oxidases ImcH and CbcL in the cytoplasmic membrane; PpcA and PpcD in the periplasm; and OmaB, OmaC, OmcB, and OmcC in the outer membrane [15–18]. Among

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these c-Cyts, OmaC, OmcB, and OmcC combined with porin-like outer membrane proteins OmbB and OmbC further form porin–cytochrome transouter membrane protein complexes [19, 20]. Based on this, electrons of the quinone and quinol pool in the cytoplasmic membrane are transferred by c-Cyts and porin-like proteins, pass the periplasm and across the outer membrane to the bacterial surface, and arrive to the electrode. Besides c-Cyts, G. sulfurreducens can also transfer electron through nanowires. These nanowires can function as electronic conduits to transfer electrons to electrodes rather than be not needed for attachment to the surface of electrodes [21]. The conductivity of Geobacter nanowires has been successfully measured through synchrotron X-ray microdiffraction and rocking-curve X-ray diffraction [22]. The measured conductivity increases with the decrease of temperature or pH. The property of temperature-dependent and pH-dependent conductivity is familiar to conductive polymers [23]. The electron transfer mechanism of nanowires is mediated by a continuous aromatic–aromatic interaction chain that is formed via the stack of aromatic amino acids of GSu PilA [22]. This mechanism is considered to be shared with metallic-like electron transfer mechanism. Actually, the nanowires are pili that are formed by protein filaments anchored in the cell envelope [23]. Previous studies have demonstrated that conductive pili is essential for G. sulfurreducens to accomplish long-range electron transport to electrodes [24]. Currently, acetogens, as anaerobes that are capable of using the Wood– Ljungdahl pathway to produce acetyl-CoA through reduction of CO2 , has been widely applied in bioelectrosynthesis for synthesis of acetate, ethanol, butyrate, and butanol [25]. Acetogens are considered to directly transport electrons to electrodes through ferredoxin that is a membrane-bound complex, which is composed of Rnf complex and Ech complex [26]. Rnf complex and Ech complex are found in Clostridium ljungdahlii and Moorella thermoacetica, respectively. Rnf complex and Ech complex participate in both electron transfer and proton transport, which can drive adenosine triphosphate (ATP) synthesis. Other acetogens such as Sporomusa ovata and Clostridium aceticum are also demonstrated to directly transport electrons to electrodes [27]. However, the mechanisms of EET of these acetogens are still unknown. 10.1.2

Indirect Electron Transfer

The mechanism by which bacteria and archaea acquire the necessary electrons for their metabolism is mainly determined by the potential of the cathode, the type of microbial catalyst, and the composition of the culture medium. IDET during electrosynthesis often indirectly involves soluble shuttles that could transport electrons from the solid electrode to the microbe. Electron shuttles are generally small redox mediators such as H2 , formate, Fe2+ , ammonia, or sulfide, which can facilitate the construction of multicarbon organics from carbon dioxide using naturally occurring carbon fixation pathways and can be electrolytically regenerated in the MES. These inorganic compounds naturally yield sufficient energy to support microbial growth and can be reduced via electrolysis. It has also been suggested that more complex soluble redox shuttles, such as flavins and DNA,

10.1 External Electron Transfer

could be produced and excreted by microbial catalyst or released in the bioelectrochemical systems (BES) electrolytes after cell death [1, 28, 29]. H2 requires only that the cathode in the MES reactor is poised at a lower potential than −0.41 V vs. the standard hydrogen electrode (SHE); therefore H2 is one of the most prevalent electron shuttles and microbial could acquire enough energy to maintain cell growth and CO2 reduction [30, 31]. If acetogens or methanogens are dominant microbial catalysts in the MES, IDET will occur via H2 , one of their preferred electron donors under autotrophic growth conditions [32, 33]. Several acetogenic bacteria including Sporomusa ovata, Sporomusa silvacetica, Sporomusa sphaeroides, Clostridium ljungdahlii, Clostridium aceticum, and Moorella thermoacetica consumed electrons from a cathode to reduce CO2 to acetate along with H2 reduction via the acetyl coenzyme A or Wood–Ljungdahl pathway [34]. Meanwhile, autotrophic methanogens could also convert CO2 to CH4 via indirect H2 mediator channels [35]. It has been reported that significant quantities of H2 are generated from the electrons coming from the cathode for the abiotic evolution and the protons migrating from the anodic chamber, which are often employed for MES applications. However, employing H2 as a redox mediator for MES is not optimal, considering its low solubility and explosive nature, which limit the usefulness of this approach for most applications. Furthermore, efficient production of H2 typically requires expensive metallic catalysts or substantial inputs of energy to overcome sluggishness in proton reduction at electrode surfaces [36]. Compared with H2 , formate is safer because of its higher solubility [10]. It will be generated electrochemically from CO2 and protons at the cathode potential poised below −1.5 mV vs. SHE [37]. This strategy has been employed to shuttle electrons from the cathode to Ralstonia eutropha for the electrosynthesis of biofuels [38]. However, one of the main disadvantages of employing a low cathode potential is that the MES system will require more electrical energy, reducing the energetic efficiency. When MES systems with a medium containing Fe3+ , iron-oxidizing bacteria, such as obligate aerobic chemolithotrophs and anaerobic photoautotrophs, can uptake electrons from electrodes along with small-molecule Fe2+ mediated. For example, Rhodopseudomonas palustris TIE-1, identified as iron-oxidizing bacteria, has been reported capable of extracting electrons by using mediated electron transfer with Fe2+ , which could be continuously regenerated at the cathode [39]. Leptospirillum ferrooxidans P3A and CF27, which use Fe2+ to respire aerobically, are also capable of obtaining electrons from electrode with Fe2+ oxidation [40]. This reduction can be accomplished with a cathode poised at a higher potential than for acetogen- or methanogen-driven MES systems, meaning that less electrical energy is required. However, thus far, electron transfer rates and chemical production rates from CO2 are significantly lower with Fe2+ oxidizing bacteria than have been observed with acetogens or methanogens in MES systems. Ammonia is a cost-efficient, abundant, safe, and soluble redox shuttle that could facilitate energy transfer to biomass. Nitrosomonas europaea, which has the ability to utilize ammonia as sole electron donor for growth, has been used as a microbial catalyst for MES in a system where ammonia could electrochemically regenerate from nitrite [41].

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Another possible IDET mechanism is that microbial catalysts could be producing and excreting their own soluble redox mediators to carry electrons from the cathode. It was recently reported that a hydrogenase and formate dehydrogenase, which are released out of cells, are adsorbed onto electrodes and mediate electron transfer between cathode and Methanococcus maripaludis [29].

10.2 Promotion of Material Development Electrode materials are not only the key components in the MFCs and microbial electrolysis cells (MECs) but also in bioelectrosynthesis. As microorganisms colonized on electrodes function as living catalysts for electron transfer, desirable microscale features of the electrodes should meet high conductivity, biocompatibility, and stability. Furthermore, the electrode provides a diffusion pathway to deliver the generated organic products and accommodate the culture media, and thus desirable macroscale feature of electrodes is porous structure. More importantly, commercialization of bioelectrosynthesis will require optimization and scaling, and thus the desired electrode materials should have low cost, excellent chemical stability, and high mechanical strength. Carbon-based materials such as graphite stick, glassy carbon, and fibrous materials (felt, paper, fibers, cloth, and foam) are the most versatile electrodes for bioconversion of CO2 in the bioelectrosynthesis systems [42]. Metal-based cathode materials such as stainless steel and nickel foam were also proved to be feasible in bioelectrosynthesis [43, 44]. Recently, there appeared numerous reports on electrode development through building a 3D porous structure and surface modification via metal or composited nanomaterials and conductive polymers (Figure 10.2) [45–48]. Good reviews of these electrode material development and electrode modification studies are available [43, 49]. Here, we summarize electrode material development and electrode modification for the improved bioelectrosynthesis.

Structure and surface design/modification

3-Dimensional

Porous

Biocompatibility

Conductivity

Increased surface area and affinitive electrode surface for microbial attachment and electron transfer

High production rate and current density

Figure 10.2 Goal and strategy of the improved cathodes for high-performance bioelectrosynthesis.

10.2 Promotion of Material Development

10.2.1

Carbon-Based Electrodes for Bioelectrosynthesis

Carbon-based electrodes have been commonly used in bioelectrosynthesis because of their good conductivity, excellent biocompatibility, low cost, and long durability, which may be important advantages for bioelectrosynthesis application [50]. Conventional carbon-based electrodes mainly consisted of graphite stick/rod, carbon cloth, carbon cloth, carbon felt, gas diffusion activated carbon (AC), etc. Graphite is a commonly used carbon material in bioelectrosynthesis because of its plane sheet structure and intrinsic merits such as relative inertness, wide electrochemical windows, high electrical conductivity, low residual current, high flexibility, reusability, recyclability, and outstanding biocompatibility. Lovley’s group firstly described the proof of concept of bioelectrosynthesis in 2010 where the cathode comprised unpolished graphite sticks, and electrosynthesis products were acetate and small amounts of 2-oxobutyrate [51]. However, the drawbacks of using graphite rod or block cathode will meet a relatively lower productivity in bioelectrosynthesis because of its limited porosity and accessible surface area for the attachment of microorganisms. However, graphite granules could provide a higher surface area compared to carbon cloth, which may be beneficial for bacterial attachment. For example, Marshall et al. investigated graphite granules as cathode in bioelectrosynthesis system incubated with brewery wastewater [52]. When cathode poised at −0.590 mV vs. SHE with CO2 as the only carbon source, the products contained methane and acetate. Additionally, Marshall et al. further prolonged incubation of brewery waste-fed electrosynthetic systems using graphite granules cathode and achieved an improved performance of electrosynthetic biocathodes [53]. Besides graphite-based cathodes, carbon cloth is also an excellent electrode material that has been widely applied in bioelectrosynthesis. For example, Ganigué et al. demonstrated for the first time transformation CO2 into butyrate in the bioelectrosynthesis system using a carbon cloth cathode [54]. This system achieved a highest butyrate concentration of 20.2 mM, with a maximum butyrate production rate of 1.82 mM/d. In addition, ethanol and butanol were also produced, which open up the potential for biofuel production in the bioelectrosynthesis system with a carbon cloth cathode. However, the carbon cloth electrodes also have a few deficiencies such as low electrocatalytic activity, high activation overpotential, high internal resistance, and rapid creation of a passivation layer on the electrode surface, which may limit their scale application in bioelectrosynthesis systems [55]. In addition, three-dimensional (3D) thin morphology carbon materials such as carbon felt are also widely used because of better electrical conductivity, flexibility, chemical stability, and higher porous and fibrous structures that permit efficient mass transfer when compared to traditional 2D graphite electrode. Furthermore, carbon cloth possessed high surface area for bacterial adherence. An early representative research by Patil et al. achieved continuous supply of reducing equivalents enabled acetate production at a rate of 19 ± 2 g/m2 d (projected cathode area) using a raw carbon felt cathodes [56]. This is a considerably high rate compared with other unmodified carbon-based cathodes. The authors found that the acetate production rate had a positive correlation to the

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projected cathode surface area, and this may be an important parameter for bioelectrosynthesis in large scale. Unlike other carbon-based electrodes, reticulated vitreous carbon (RVC) electrode composed solely of vitreous carbon with a honeycomb and open-pore foam structure [57]. It possesses low resistance to fluid flow, strong chemical and heat resistance, high void volume, and excellent conductivity. Furthermore, the high surface area-to-volume ratio of the RVC represent a typical characterization of three-dimensional (3D) electrodes, could provide a large surface area, and thus increase the active biomass and current consumption per given volume unit [7]. The pore size (number of pores per linear inch [ppi]) was used as a parameter to characterize the structure of a particular sample of RVC. The commonly used pore sizes varied from 10 to 100 ppi in bioelectrosynthesis. However, RVC was demonstrated not to have the favorable surface characteristics for bacterial adherence and electron transfer [58]. Therefore, RVC electrodes are often modified with carbon nanotube-based materials and applied in bioelectrosynthesis [7], which will be described in the later section. 10.2.2

Metal-Based Electrode for Bioelectrosynthesis

During the development of bioelectrochemical technology, metal-based materials played a key role in the preparation of both anode and cathode. Previous studies have reported that stainless steel and nickel could act as cathode for bioelectrosynthesis because of their superior mechanical properties, electrical conductivity, and corrosion resistance. In this regard, stainless steel material has been commonly used either as cathodes or integrated with the carbon-based materials in bioelectrosynthesis systems. For example, Dumas et al. investigated stainless steel cathode polarized at −0.60 V vs. Ag/AgCl in bioelectrosynthesis reactors inoculated with G. sulfurreducens [59]. Results showed that fumarate was reduced to succinate when the current was around 20.5 A/m2 . H2 could mediate the electron transfer in CO2 reduction conducted by bacteria at the cathode; however, the existence of overpotential of electrode material limited H2 evolution. Stainless steel holds the ability to facilitate H2 evolution especially in the presence of weak acids, which might be necessary for the effective CO2 reduction in bioelectrosynthesis. These were demonstrated by Bajracharya et al. [43]. In their study, CO2 reduction in bioelectrosynthesis was investigated at hydrogen evolving potentials, separately by a mixed culture and C. ljungdahlii, using a stainless steel and graphite felt assembly as cathode. The bioelectrosynthesis reactor with mixed culture produced acetate at the maximum rate of 1.3 mM/d, accompanied by methane and hydrogen at −1.1 V vs. Ag/AgCl. However, the generated H2 bubbles from the stainless steel cathodes might impact the biofilm formation [60]. This question deserves further study in bioelectrosynthesis system. Until very recently, nickel foam has attracted attention in bioelectrosynthesis because of its unique 3D scaffold structure with a high surface area [44]. A macropore size structure of nickel foam is beneficial for bacterial colonization and also allows both the mass transfer of carbon dioxide and all nutrients and protons to these bacteria. For example, Song et al. reported that an acetate production rate of 0.102 g/l d was achieved through 28 days in bioelectrosynthesis reactor

10.3 Modified Electrodes for High Bioelectrosynthesis

with the nickel foam cathode [44]. More importantly, nickel foam was relatively cheap and commercially available when acted as supported materials for electrode modification. However, the intrinsic lower electrochemical activity might limit the bioelectrosynthesis performance. Therefore, more studies have focused on nickel foam electrode modification with carbon or metal materials, which will be described in the following section.

10.3 Modified Electrodes for High Bioelectrosynthesis The advances in anode materials in MFC revealed a porous surface of the electrodes, which leads to a higher specific surface area to interface with bacteria and improved electron exchange on the electrode surface. Based on this design concept, similar electrodes were also used in bioelectrosynthesis. Meanwhile, in the development of immobilization technology, especially materials science, further breakthroughs have been obtained in the design of electrodes throughout the 1980s and 1990s [61]. Therefore, a variety of modified electrodes were fabricated for the enhanced bioelectrosynthesis. It needs to be emphasized that electrode will undergo pretreatment such as chemical or thermal treatment to remove impurities from its surface before electrode surface modification. This could potentially expose the reactive surfaces for better decoration and immobilization. It has been extensively recognized that an increased electrode surface may contribute to the mass production of acetate, methane, and other highly valuable products. In this respect, a three-dimensional electrode is more attractive than a two-dimensional one because of its extensive interfacial electrode surface and high mass transfer. Consideration of dimensionality and function of electrodes, we firstly summarize design of porous electrodes modified with carbon-based and metal-based materials and then discuss functional electrodes for bioelectrosynthesis. 10.3.1

Electrode Modification with Carbon-Based Materials

Carbon-based materials can not only be used as substrate electrodes because of their high conductivity, stability, biocompatibility, and low cost but also be served as modification materials for electrodes in bioelectrosynthesis system. A considerable improvement in electrode modification-based acetate production rate has been obtained by using purposely modified electrode with carbon-based materials in bioelectrosynthesis systems. Some attractive carbon-based materials that resulted in enhancement in bioelectrosynthesis performance indicators are briefly reviewed below. 10.3.1.1

Carbon Nanotube

Carbon nanotubes (CNTs) as one of most commonly used carbon-based materials have drawn remarkable attention in bioelectrodes because of their unique dimensions and structure-sensitive properties [62]. CNTs consist of one or more layers of graphene (denoted as single-walled nanotube, SWNT, or multiwalled

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nanotube, MWNT), with open or closed ends [63]. The preferable electrical conductivity allows the utilization of CNTs for electrode modification and in combination with high electrocatalytic activity could mediate electron transfer among electrodes and bacteria, which offers great promise for bioelectrosynthesis. For example, Jourdin et al. reported a novel a novel, biocompatible, highly conductive, three-dimensional cathode fabricated by direct growth of flexible multiwalled carbon nanotube (MWCNT) on RVC (NanoWeb-RVC) for the enhancement of bioelectrosynthesis [7]. The acetate production rate (1.3 mM/cm2 d) and current density (3.7 mA/cm) were 2.6- and 1.7-fold higher, respectively, when compared to a carbon plate control cathode for the microbial reduction of carbon dioxide by mixed cultures. They found that the porous and three-dimensional structure of electrode contributed to improved bioelectrosynthesis. In addition, the directly grown CNTs on RVC surfaces may create a high density of active electron transfer sites, which could then interact with the microbial cells growing on top. Subsequent studies by Jourdin et al. further demonstrated that using RVC electrodes with macropore sizes of about 0.6 mm in diameter modified with MWCNT was found to be optimal for achieving a good balance between total surface area available for biofilm formation and effective mass transfer between the bulk liquid and the electrode and biofilm surface [64]. Additionally, Zhang et al. reported cotton and polyester textile fiber cathodes modified with SWNTs for bioelectrosynthesis, and the production rate of acetate inoculated with acetogenic bacteria, i.e. S. ovata, was 3.4- and 3.2-fold higher than the above SWNT-modified cathodes when compared to control (Figure 10.3a) [8]. However, it has been reported that SWNTs have a cellular toxicity that could lead to proliferation inhibition and cell death [65]. Furthermore, SWNT showed less toxicity to mature biofilms compared to bacterial cells in other biofilm phases, which may have a negative effect on biofilm formation of cathode in the initial stage of bioelectrosynthesis operation. Therefore, the amount of CNTs on cathodes should be controlled at a security level, and this work deserves further study in the future. 10.3.1.2

Graphene

Since isolation in 2004, graphene has been considered as an emerging material with promising application potentials in many fields, such as supercapacitors, sensors, and energy storage because of its unique nanostructure and extraordinary properties such as high surface area, excellent conductivity, and extraordinary electrocatalytic activities, etc. [66, 67]. Graphene has a two-dimensional (2D) structure with free-standing carbon atoms packed into a dense honeycomb crystal structure and usually prepared through thermal annealing of graphene oxide (GO), which is synthesized via a modified Hummers’ synthesis [68]. Such synthesized graphene is called chemically modified graphene or reduced graphene oxide (rGO) possibly because it is more defective than pristine graphene [69]. Actually, graphene-based materials have been widely used for both anode and cathode modification in BES for improvement of power generation through establishment of porous electrodes with 3D structure [70, 71]. This electrode

10.3 Modified Electrodes for High Bioelectrosynthesis

(a)

(b)

(c)

(d)

(e)

(f)

Figure 10.3 SEM images of different material-modified cathodes. (a) CNT-modified cotton cathode. Source: Zhang et al. 2013 [8]. Reproduced with permission of Royal Society of Chemistry. (b) 3D graphene-coated carbon felt cathode. Source: Aryal et al. 2016 [49]. Reproduced with permission of Elsevier. (c) Graphene-coated nickel foam cathode. Source: Song et al. 2017 [44]. Reproduced with permission of John Wiley & Sons. (d) Nickel nanowire on nickel nanoparticle-coated graphite cathode. Source: Nie et al. 2013 [47]. Reproduced with permission of Royal Society of Chemistry. (e) PANi–PAN-modified carbon cloth cathode. Source: Zhang et al. 2013 [8]. Reproduced with permission of Royal Society of Chemistry. (f ) Chitosan-modified carbon cloth cathode. Source: Zhang et al. 2013 [8]. Reproduced with permission of Royal Society of Chemistry.

modification approach with graphene applies equally to bioelectrosynthesis process for enhancement of acetate production. For example, Aryal et al. reported a 3D graphene-functionalized carbon felt composite cathode for bioelectrosynthesis [49]. The scanning electron microscope (SEM) of the modified cathodes is shown in Figure 10.3b. They revealed that the specific surface area measured with the Brunauer–Emmett–Teller (BET) method was 2.2-fold higher with 3D

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graphene electrode (2.99 m2 /g) than plain carbon felt electrode (1.36 m2 /g). It should be pointed out that BET surface area, electrochemically active surface, and biofilm-covered area are commonly used to analyze the performance of electrode materials in bioelectrochemical process. The BET surface is relatively convenient to report the performance of the 3D graphene functionalized carbon felt in bioelectrosynthesis systems. The acetate production rate was increased by 6.8 folds in bioelectrosynthesis systems with 3D graphene modified cathode, where the cathode inoculated with S. ovata. Previous literature studies have demonstrated that graphene could not only create more electroactive sites for microbe colonization that resulted in an improved electron exchange capacity but also improve electron transfer from cathode to a microbial catalyst [70]. In addition, a higher BET-active surface area of electrodes will promote the current density and further leads to a decrease of ohmic loss [72]. This maybe another reason for the improved acetate production rate in the above bioelectrosynthesis system using graphene-modified carbon felt. In this case, cathode modification with graphene results in manifold improvement in the bioelectrosynthesis productivity. Typically, graphene as a good carbon support can provide edge plane anchor sites for metal catalyst. Previous studies had demonstrated that the heterogeneity of graphene with edge planes can better stabilize and enhance the catalytic activity of the metal catalyst [73]. Graphene has been employed as support for platinum nanoparticles for cathode decoration in proton exchange membrane fuel cell (PEMFC) [74]. This cathode showed superior polarization performance when compared to commercial carbon black-supported Pt cathode. Based on unique properties of graphene, a combination of graphene with metal is also an alternative strategy for cathodes functioned with both three-dimensional porous and catalytic ability in bioelectrosynthesis. For example, Song et al. fabricated a novel three-dimensional graphene–nickel foam electrode using hydrothermal approach for bioelectrosynthesis to improve acetate production (Figure 10.3c) [44]. The acetate production rate was 1.8 times higher in the bioelectrosynthesis reactor using the graphene–nickel foam cathode (3.11 mM/d) by a mix culture compared to untreated nickel foam (1.73 mM/d). Theoretically, the produced acetate from the cathode would decrease the pH of catholyte while carbon dioxide was continuously bubbled into the cathode chamber, which neutralized the pH in the bioelectrosynthesis reactor using graphene–nickel foam cathode. However, the nickel foam experienced slow corrosion under this neutral condition, which reverse disturbed the pH of catholyte. This partially resulted in the current of the bioelectrosynthesis using nickel foam cathode and further weakened the bioelectrosynthesis performance. Furthermore, this also proved that the graphene coating could effectively reduce the corrosive effect and stabilize the solution pH and current. However, the potential disadvantage when employing reduced graphene in cathode modification is that it is a negatively charged material, which would repulse a negatively charged bacterial cell [75]. Recently, tetraethylene pentamine (TEPA) as a positively charged molecule containing amino groups is used to functionalize rGO, which may be beneficial for bacterial attachment [76]. Chen et al. fabricated a rGO–TEPA-modified cathode for bioelectrosynthesis

10.3 Modified Electrodes for High Bioelectrosynthesis

system [46]. Results showed that electroactive biofilms with a unique spatial arrangement were formed with S. ovata cells, resulting in a bioelectrosynthesis process more performant with rGO–TEPA-modified cathodes and the acetate production rate (0.9 ± 0.15 mM/d) was increased by 3.6-fold with the formation of dense biofilms on modified cathodes (0.25 ± 0.02 mM/d). 10.3.1.3

Activated Carbon

Activated carbon (AC) is a common carbon-based material that has a large surface area in excess of 500 m2 /g, and this high surface area allows high extent of bacterial attachment and further increases the surface area available for electron exchange between bacteria and the electrode [77]. AC is usually facilely and economically produced from various carbonaceous sources such as nutshells, wood, coconut husk, coal, etc.; its low cost is very competitive compared to CNT, graphene, and other carbon-based materials. In MFCs, AC often acts as anode supporting materials for the improved power generation because of its high specific area, moderate electrical conductivity, and good biocompatibility [78]. In addition, AC can also serve as a catalyst for oxygen reduction reaction (ORR) on the cathode of MFCs [79]. It has been reported that AC can also be used as a redox mediator because it contains different functional groups (active sites) capable of being reduced and oxidized, which is beneficial for electron transfer between bacteria and electrodes [80]. As such, AC is expected to fit in bioelectrosynthesis. For example, Bajracharya et al. fabricated a hydrophobic gas diffusion layer cathode by the aid of AC for bioelectrosynthesis, which creates a three-phase interface at the electrode [81]. When the cathode potential is controlled at −1.1 V vs. Ag/AgCl, bioelectrochemical CO2 reduction produced acetate, ethanol, and butyrate. The maximum acetate production rate achieved 238 mg/l d for 20% CO2 gas mixture. 10.3.2

Electrode Decoration with Metal-Based Materials

Metal-based materials in bioelectrosynthesis can serve not only directly as cathode but also as decorated materials for electrode promotion. Usually, metal-based materials refer to nanosized materials that are applied in electrode decoration. These metal nanomaterials have been widely investigated for electrode design in bioelectrochemical system because of their unique physical and chemical properties. Metal nanomaterials consist of metal and metal oxide nanoparticles or nanowires. Owing to their low cost, favorable biocompatibility, high electric conductivity, and catalytic activity, metal and metal oxide nanomaterial decoration has been demonstrated to be effective to increase the surface area of electrode and accelerate the electron transfer among bacteria and electrode in bioelectrosynthesis process. In recent years, researchers have prepared cathodes decorated with various metal nanomaterials such as gold in order to improve bioelectrosynthesis performance. Zhang et al. systematically explored gold, palladium, and nickel nanoparticle-decorated cathodes in bioelectrosynthesis reactors inoculated with S. ovata [8]. They found that the acetate production rate was 6-, 4.7-, and 4.5-fold faster for gold, palladium, and nickel decorated cathodes, respectively,

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when compared to untreated control cathode. This was attributed to the high conductivity and the low charge transfer resistances of the metal nanoparticles. In comparison to nanoparticle (0D), the nanowires (1D) take advantages of unique properties such as porous structure, higher surface area, and conductivity, which may therefore provide a more effective local contact with the bacteria of cathodes. In a recent study, Nie et al. developed a novel cathode decorated with nickel nanowires anchored to graphite for the enhancement of bioelectrosynthesis performance (Figure 10.3d) [47]. The interfacial area of cathode and interfacial interactions between the cathode surface and the microbial biofilm were significantly increased through decoration by nickel nanowires. Acetate production rate was about 282 mM/d m2 with nickel nanowire-decorated cathode in bioelectrosynthesis reactor by Sporomusa with, which was increased by 2.3-fold than that with uncoated graphite cathode. These results also revealed that nickel nanowire-decorated graphite afforded enhanced microbial-catalyzed reduction of CO2 , due in large part to the superior surface topography resulting from the nature of the porous nanowire layer. Besides metal nanoparticles, metal oxides such as Fe2 O3 have been proved to be attractive nanomaterials for electrode decoration in bioelectrochemical systems [82]. Fe2 O3 nanoparticles possess a semiconductive property that may lead to better microbial EET. For example, Cui et al. prepared a 3D hierarchically porous electrode through in site microwave pyrolysis of ferrocene on carbon felt electrode (i.e. Fe2 O3 -decorated electrode) [45]. They demonstrated that acetate production rate was 2.48 × 105 mM/d m3 in a bioelectrosynthesis reactor using Fe2 O3 -decorated carbon felt cathode, which shows fivefold increase compared to plain carbon felt cathode (5.19 × 104 mM/d m3 ). This Fe2 O3 nanoparticle on carbon felt electrode provides a higher specific surface area that also helps to improve bioelectrosynthesis performance. In another study, titanium dioxide (TiO2 ) is used to modify electrode for improving bioelectrosynthesis performance [83]. A TiO2 and silicon (Si) mixture-decorated electrode was prepared and used as a cathode for the photosynthesis–bioelectrosynthesis hybrid system. In this system, Si–TiO2 nanowire arrays on the electrode are used as the light-capturing units to mimic the “Z-scheme” and S. ovata as the cellular catalyst, which can effectively reduce CO2 and produce acetate under simulated sunlight. This photosynthesis–bioelectrosynthesis hybrid system with Si–TiO2 nanowire electrode showed a high reaction rate of CO2 reduction, and the nanowire array on electrodes provides a local anaerobic environment that allows strict anaerobes to continue CO2 reduction aerobically (21% O2 ), which may be important for practical application in the future. 10.3.3

Electrode Decoration with Other Materials

In recent years, small molecular and polymeric materials have attracted increasing attention from researchers for their advantages of readily tunable bandgaps, comparable biocompatibility, excellent electroactivity, ease of modification, and good processability over conventional inorganic nanomaterials [84]. With these characteristics, small molecular and polymeric materials such as chitosan, polyaniline, polypyrrole, polyvinyl alcohol, Nafion, and

10.3 Modified Electrodes for High Bioelectrosynthesis

3-aminopropyltriethoxysilane are widely used for electrode decoration in bioelectrochemical systems. Furthermore, these polymeric materials can also function as receptors in the form of a polymer matrix or mediators, which is beneficial for adherent of microorganisms and electron transfer. It is apparent that if electron transfer is guaranteed, the catalytic reaction of enzyme (isolated or supported by microbials) will be sustainable, which is the reason that conducting electroactive biocompatible polymers are served as materials for immobilization, especially when applied in electrode decoration for bioelectrosynthesis. Polyaniline (PANi) as a conjugated polymer has been regarded as a promising battery material because of its environmental stability, excellent conductivity and biocompatibility, and interesting redox behavior. Previous studies revealed that the electrochemical capacitance of PANi is more than 400 F/g because of the existence of redox states in the presence of dopants [85]. Because being synthesized for the first time over two decades ago, various methodologies have been demonstrated to be effective in the preparation of nanostructured PANi in the form of nanofibers, nanowires, and nanotubes. Electrospinning is a novel technique for developing 3D scaffolds of PANi with high surface-to-volume ratios, microscale porosity, and fiber interconnectivity. The electrospinning synthesized PANi fibers provide a nanofibrous scaffold with strong adsorbability and enough space for biomacromolecules [86]. With these advantages, PANi showed great potential for application in the preparation of electrodes in bioelectrochemical systems especially in bioelectrosynthesis. For example, Zhang et al. developed a PANi–polyacrylonitrile (PAN)-decorated carbon cloth cathode using electrospinning technique for the bioconversion of CO2 in a bioelectrosynthesis system with S. ovata (Figure 10.3e) [8]. The production rate of acetate was threefold higher with PANi–PAN cathode than that with control carbon cloth cathode, where the recovery of electrons consumed in acetate production was 85 ± 7%. The microporous structure of PANi–PAN was suggested to enhance the efficiency of EET in this case. Chitosan (CHI) is a unique physicochemical biopolymer that is most extensively used for sensor application because of its good adhesion and mechanical strength, excellent biocompatibility, and nontoxicity. CHI has numerous reactive amine groups that play a significant role in dissolubility. The pK a value of CHI is about 6.3 that results from the primary amino groups. Because the pH is above the pK a , the amino groups of CHI will be deprotonated and further leads to an insoluble status. Previous studies have demonstrated that CHI hydrogel can be attached to the electrode and retain its natural properties. In addition, CHI can be embedded on the surface of carbon cloth through the reaction between –COOH groups on the electrode surface and –NH2 groups on the chitosan. CHI is a cationic polysaccharide owing to protonation of amino groups and thus possesses positive charge. As already known, gram-negative microorganisms such as S. ovata and G. sulfurreducens usually show a negative outer-surface charge. Typically, an untreated carbon cloth has a neutral charge. Therefore, decoration of carbon cloth electrode with positive-charged CHI improves the electronic interaction between the cells and the electrodes, which will enhance bioelectrosynthesis performance, which has been proved by Zhang et al. [8] (Figure 10.3f ). The production rate of acetate in bioelectrosynthesis using CHI-decorated

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cathode was 7.6-fold higher than that with plain carbon cloth cathode. S. ovata is a well-studied electroautotroph in bioelectrosynthesis system, and it shows poor biofilm growth on the raw carbon cloth, carbon felt, or graphite cathodes. However, confocal laser-scanning fluorescence microscopy results showed that the cell density of CHI-decorated cathodes was 3.02 ± 1.95 × 107 cells/cm2 in bioelectrosynthesis system, which was much higher than that with unmodified carbon cloth cathode (3.98 ± 1.24 × 106 cells/cm2 ). Scanning electron microscopy also demonstrated that a thin layer of chitosan is coated on the surface of carbon cloth, where the formed pore sizes are suitable for microbial access. All these may be the possible reasons for the increase of acetate production rate. Besides PANi and CHI, other polymeric materials such as cyanuric chloride and 3-aminopropyltriethoxysilane are also used for electrode decoration in bioelectrosynthesis because of their nature of positive charge. Acetate production rate in bioelectrosynthesis system with S. ovata using cyanuric chloride and 3-aminopropyltriethoxysilane-decorated cathodes was 6.8- and 3.0-fold higher than that with control cathode, respectively [8]. However, not all electrode decoration methods with positive charge treatment at the surface are sufficient in improving bioelectrosynthesis performance. For example, treatment of carbon cloth electrode with ammonia gas can enhance power generation in MFCs but not stimulate acetate electrosynthesis [87].

10.4 Interspecies Electron Transfer Pathway Fermentative acidogenic bacteria often collaborate with other methanogenic archaea to completely converse organic matter to methane, which is termed syntrophy. During this syntrophic process, fermenting bacteria produce surplus reducing power that should be disposed, and while methanogens are the sink of electrons disposed from organic matter oxidizers. Electron transfer among these different microorganisms is called interspecies electron transfer. The phenomenon of interspecies electron transfer is usually occurred in microbial fermentation for methane formation under anaerobic environments. Until now, there are three mechanisms of interspecies electron transfer that have been proposed, i.e. interspecies H2 transfer, interspecies formate transfer, and direct interspecies electron transfer. Interestingly, interspecies electron transfer also happens in bioelectrosynthesis process. However, few studies focus on interspecies formate transfer in bioelectrosynthesis except interspecies H2 transfer. Direct interspecies electron transfer has been discussed in the former section (see Section 5.1). The mechanism of interspecies H2 transfer is firstly developed by Omelianski in 1830s, where he found that the microorganisms are related to the production of H2 , acetate, and butyrate and further demonstrated that the microorganism facilitated the reaction between H2 and CO2 , resulting in the production of methane. This microorganism was isolated and named Methanobacillus omelianskii, which lives on acetate from oxidation of ethanol in microbial catabolism process and produced electrons are used to reduce CO2 and/or simultaneously oxidate H2 to

10.5 Future Perspectives

methane. This mechanism is further developed by other researchers until Wolin proposed a complete mechanism of interspecies H2 transfer in 1982 [88]. Electron transfer rate from cathode is often regarded as a rate-limiting step during bioelectrosynthesis process on a commercial scale [89]. Usually, an overpotential of more than 200 mV (vs. SHE) is applied to achieve cathodic electron uptake for the production of acetate or methane. To address this limitation, researchers have attempted various approaches to achieve effective electron transfer and thus improve the bioelectrosynthesis rate. Recently, Deutzmann and Spormann demonstrated that interspecies H2 transfer may be an alternative way for the enhancement of bioelectrosynthesis [90]. They studied that a defined coculture contained a Fe(0)-corroding strain IS4 to catalyze the electron uptake reaction from cathodes forming H2 as the intermediate. M. maripaludis and Acetobacterium woodii were served as cocultures for hydrogenotrophic production of methane and acetate, respectively. Results showed that the cocultures of IS4 and M. maripaludis accomplished methanate production rates of 0.1–0.14 μmol/cm2 h at −400 mV vs. SHE and 0.6–0.9 μmol/cm2 h at −500 mV. The rate of acetate is 0.21–0.23 μmol/cm2 h at −400 mV and 0.57–0.74 μmol/cm2 h at −500 mV for the cocultures of IS4 and A. woodii.

10.5 Future Perspectives In summary, we have reviewed numerous materials for electrode modification to improve bioelectrosynthesis performance. However, there still exist many great challenges that hamper further advances in design of electrodes in bioelectrosynthesis system. High surface area and biocompatibility are two main parameters that have been taken into consideration for building electrodes in bioelectrosynthesis system. Hence, electrode with 3D structure remarkably increases the surface area. Besides biocompatibility and high surface area, other parameters such as electric conductivity may also play important roles in bioelectrochemical CO2 reduction process. Additionally, the potential of electrode should be seriously considered in bioelectrosynthesis system. As the threshold potential of cathode for hydrogen evolution is 0.6 V vs. Ag/AgCl under biological conditions, however, the threshold potential will negatively shift to a certain extent owing to the electrode overpotentials. Although microorganisms of cathode bacteria facilitate to reduce CO2 at less negative potential than that required for hydrogen evolution, direct electron uptake from electrode could potentially occur [43]. Therefore, the threshold potential of CO2 reduction should be taken into consideration when developing electrodes in bioelectrosynthesis systems. One of the main challenges for bioelectrosynthesis process is the cost of electrode materials, which will limit the practical application of bioelectrosynthesis technology in industry market. Although acetate production efficiency has been considerably increased in lab-scale reactors by the use of novelty electrode materials, pilot-scale production acetate has still not been investigated. Bioelectrosynthesis reactors with graphene and noble metal nanoparticle-modified cathodes have exhibited excellent performance, but these materials are too expensive to

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be amplified in pilot or large scale. Previous studies have revealed that the cost of acetate produced from bioelectrosynthesis systems is much more expensive than that with other biological technologies because of the extra costs accompanied by the use of electrode materials and ion exchange membranes. In this case, the development of electrodes must combine high performance and low cost in bioelectrosynthesis in the future. Meanwhile, successful development of electrode in bioelectrosynthesis requires joint effort from different disciplines, biology, physics, engineering science, and material science. In addition, the achieved progress in electrode design for other microbial electrochemical technologies can also be investigated for electrode design for bioelectrosynthesis system. Focusing on the recent advances in improving bioelectrosynthesis process based on electrode development, the future of this technology is bright. Bioelectrosynthesis may be an important technology in the fields of biomanufacture in the future, although the efficiency of bioelectrosynthesis is relatively low. The core of bioelectrosynthesis is electron transfer process from cathode to cells; however, key knowledge gaps remain in our understanding of this transfer process. It remains unclear how many microorganisms can transfer electrons directly in nature? The mechanisms of how microorganisms interact with cathodes are not well understood. Therefore, bioelectrosynthesis performance may be improved after these electron transfer mechanisms are well studied.

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11 External Electron Transfer: Pathway, Mechanism, and Microorganisms Involved Cong Huang 1 , Jun Nan 1 , and Aijie Wang 1,2 1 Harbin Institute of Technology, State Key Laboratory of Urban Water Resource and Environment, No.73. Huanghe Road, Nangang, Harbin, 150090, China 2 Chinese Academy of Sciences, Key Laboratory of Environmental Biotechnology, Research Center for Eco-Environmental Sciences, No.18. Shuangqing Road, Haidian, Beijing, 100085, China

11.1 External Electron Transfer of Cathode Electroactive microorganisms (EAMs) are a class of microorganisms that can transfer electrons with extracellular solid-state carrier electrodes. Some EAMs can transmit electrons to electrodes to generate electric current, and some EAMs can absorb electrons from electrodes to obtain target products. Some EAMs have the ability to output and absorb extracellular electrons [1, 2]. At present, the most studied EAMs that can absorb extracellular electrons are the genus Clostridium, Sporomusa, Geobacter, and Shewanella [3, 4], which synthesize CO2 into acetic acid, butyric acid, ethanol, etc., and secrete it to the extracellular, while obtaining the microorganism itself. EAMs can absorb extracellular electrons and reduce U(VI), nitrate, and other pollutants into U(IV), nitrite, N2 , and other nontoxic substances [5]. In addition, EAMs can also absorb the extracellular electrons needed to synthesize the organic matter required for their own growth. Acidithiobacillus Geobacter are the first EAMs that have been shown to utilize the cathode electrons to synthesize intracellular organics for self-growth [6]. Several strains of the genus Geobacter and Shewanella, G. metallireducens GS-15, G. sulfurreducens PCA, and S. oneidensis MR-1 can both export electrons and absorb extracellular electrons and could transmit electrons in both directions. G. sulfurreducens had different gene expression profiles when exporting and absorbing electrons. If knocking out the omcZ and pilA genes of G. sulfurreducens, which are crucial in the export electron process, there is no effect on the absorption of extracellular electrons, indicating that the electron transport paths in these two directions are not simple, reversal, and mobile [7]. It also indicates that the study of the mechanism of absorption of extracellular electrons is a new field compared to the study of EAMs output electrons. Electron-driven intracellular reduction (synthesis) metabolism on the cathode microbial absorption electrode is a key process of bioelectrosynthesis, Bioelectrosynthesis: Principles and Technologies for Value-Added Products, First Edition. Edited by Aijie Wang, Wenzong Liu, Bo Zhang, and Weiwei Cai. © 2020 Wiley-VCH Verlag GmbH & Co. KGaA. Published 2020 by Wiley-VCH Verlag GmbH & Co. KGaA.

11 External Electron Transfer: Pathway, Mechanism, and Microorganisms Involved

Indirect electron transfer (lET) Substrate e– Electrode (electron donor)

282

Figure 11.1 Mechanisms of external electron transfer.

Mediators Cell

Mediators (e.g. methyl viologen, neutral red, flavins, H2, formic acid)

Direct electron transfer (DET) e–

Cell Direct contact (via redox outer-membrane proteins)

and external electron transfer (EET) rate is one of the decisive factors of bioelectrosynthesis efficiency. Therefore, the analysis of bioelectrosynthesis cathode EET mechanism is the basis for improving the electron transfer rate of the cathode microorganisms and achieving high yield of bioelectrosynthesis. At present, the known EET processes of cathodic EAMs can be divided into two categories [8]: indirect delivery based on electron mediators and direct delivery based on membrane-bound proteins (Figure 11.1). 11.1.1

Interspecies Electron Transfer (IET)

Interspecies electron transfer (IET) is an electron transfer process between microorganisms and electrodes by exogenous or endogenous redox-active small molecular substances as reversible electron mediators. Synthetic exogenous electron mediators such as neutral red and methyl viologen are shown to mediate extracellular electron transport in cathodic microbes. The yield of butyric acid produced was increased 1.3 times when methyl viologen or neutral red were added into the microbial electrosynthesis (MES) by inoculated Clostridium tyrobutyricum BAS7 [9]. Neutral red can significantly increase the coulombic efficiency of volatile fatty acids produced in the process of MES cathode reduction of CO [10]. The introduction of synthetic exogenous electron mediators significantly increases the rate of electron exchange or regulates intracellular reducing power (Nicotinamide adenine dinucleotide [NADH] levels) by changing the direction of electron flow [11]. A variety of intermediate metabolites such as H2 , formic acid, and ammonia produced by the MES cathode can be used as endogenous electron mediators [10, 12]. Blanchet et al. [13] compared the reduction of CO2 at different electrode potentials of cathodes inoculated with activated sludge. It was found that the yield of acetic acid reached 244 ± 20 mg/l at an electrode potential of −0.66 V (vs. SHE [standard hydrogen electrode]) and no production of acetic acid was detected at −0.36 V. Although the two potentials are thermodynamically favorable for the reduction of CO2 to acetic acid (E0 HCO−3 /acetic acid = −0.28 V),

11.1 External Electron Transfer of Cathode

H2 production is only present on the electrode when the potential is lower than −0.41 V, indicating that H2 plays a key role in the absorption of extracellular electrons in the cathode microorganism. The H2 -driven extracellular electron transport process produced by the cathode has been reported in several MESs [14, 15]. It has been found that electrode biofilms may decrease the overpotential of hydrogen evolution in the cathode; therefore, the EET rate was accelerated by the H2 -mediated process [14]. When the MES cathode potential is between −0.5 and −0.6 V, methanogens and acetogens can use H2 and formic acid as electron mediators. Hydrogenase and formate dehydrogenase produced by the release of microbial cells to the extracellular or cell cleavage at the cathode electrode interface can catalyze the formation of H2 and formic acid [16]. 11.1.2

Direct Electron Transfer (DET)

Respiration near the surface of the electrode-producing flora tends to transfer electrons using the direct electron transfer (DET) mechanism because the enzyme of c-type cytochromes (c-cyts) exposed to their outer cell membrane is in direct contact with the electrode surface. Many electrogenic microflora mediate intracellular electrons to the surface of the electrode via the DET mechanism. DET of microorganisms attached to the electrodes do not need to rely on any electron mediator to achieve electron exchange by direct contact of the redox proteins in the outer membrane or periplasm with the electrodes. The studies on the DET mechanism of cathodic EAMs have focused on G. sulfurreducens and Shewanella oneidensis, which are model strains for studying EET mechanisms. G. sulfurreducens and S. oneidensis have been shown to have DET properties (usually the intracellular electrons of the electrogenic microbial output are referred to as positive EET, and the absorption of extracellular electrons by electrophilic microorganisms is called reverse EET). Strycharz et al. [7] showed that the G. sulfurreducens PCA strain under the condition of electron absorption, the transcriptional abundance of the PccH protein gene encoding G. sulfurreducens PCA was relative higher, and the removal of the protein gene severely inhibited its function of absorbing extracellular electrons. Santos et al. [17] obtained the PccH protein of G. sulfurreducens by graphite electrode and studied the thermodynamic and kinetic properties of PccH protein by cyclic voltammetry, which further proved that PccH protein plays a key role in the absorption of extracellular electrons. Ross et al. [18] gave direct evidence for the electron-withdrawing mechanism of S. oneidensis by knocking out related protein genes. They found that deleting the genes for FccA, MtrA, MtrB, and CymA proteins in S. oneidensis MR-1 inhibited the electron transport from the electrode to the microbe, suggesting that the Mtr pathway plays a key role in absorbing electrons. It is inferred that the transmission path when MR-1 absorbs electrons is electrode → MtrC → MtrB → MtrA → CymA → menadione cycle in the intima → FccA in the periplasm. Acetogens and methanogens have been reported to absorb the electrons of acetic acid or methane at a relatively high potential (−0.4 to −0.5 V), which is not conducive to the generation of H2 as an electron mediator under the potential conditions. They may also have direct EET mechanisms [19, 20], but they are still controversial and require further study

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of their molecular mechanisms. In addition, electrogenic microorganisms can directly export intracellular electrons and achieve longer range electron transfer by means of pili called “bacterial nanowires” [21]. Recently, some conductive materials have been tested to transport electrons in a manner that compensates for the conduction of nanowires. It has shown that magnetite can promote microbial extracellular electron transport. A mutant strain of G. sulfurreducens was used to prove the proof of OmcS deficiency. The results show that magnetite compensates for the extracellular electron transfer redox of Fe(III) in OmcS-deficient strains [21]. In G. sulfurreducens PCA, a group of proteins have been suggested to mediate the transfer of electrons across the outer membrane, the transouter membrane porin-cytochrome (Pcc) protein complex commonly found in Gram-negative bacteria [8]. Recently, some conductive materials have been tested to transport electrons in a manner that compensates for the conduction of nanowires. Further investigation revealed that magnetite attached to the electrically conductive pili of Geobacter species in a manner reminiscent of the association of the multi-heme c-type cytochrome OmcS with the pili of G. sulfurreducens [8]. The monoheme c-type cytochrome (Pcc) was shown be a common protein complex that mediates the transfer of electrons to the outer membrane in G. sulfurreducens PCA [22]. Ishii et al. [6] combined the electrochemical method and chemical labeling method to identify the electron transport path of the carbon assimilation process of A. ferrooxidans. When Fe2+ was used as an electron donor, the transfer of electrons from the outer membrane to the inner membrane of A. ferrooxidans is divided into two routes: “energy absorption” and “discharge.” In this study, the microorganisms on the electrodes in the current-stable MES were first sterilized by in situ deep ultraviolet (254 nm), and the current was inhibited, indicating that the current mainly consisted of A. ferrooxidans attached to the electrodes as electron acceptors. Direct electron transport based on direct electrode-to-cell contact is generally considered to have faster kinetics and lower energy loss than indirect electron transfer [23]. Limited device and electrode sizes in bioelectrosynthesis systems often fail to provide a sufficiently large contact area, limiting the utilization of DET. In addition, in thicker electrode biofilms, the outer microbial cells usually have higher metabolic activity than the cells encapsulated inside the biofilm, and the direct electron transport pathway does not seem to guarantee the rapid exchange of electrons. The rate of indirect electron transfer often depends on the concentration of the electron mediator, so that a rapid EET can be achieved by increasing the amount of electron mediator in the MES electrolyte.

11.2 Promotion of Material Development Surface modification of electrode materials is the most common material development method. In general, the methods for surface modification of electrode materials are summarized as follows: (i) pretreatment such as chemical immersion or heating [23], (ii) polymer-grafted electrode surface with high

Electrode (electron donor)

11.2 Promotion of Material Development

(a)

(b)

(c)

(d)

Figure 11.2 The method of material promotion. (a) Unmodified electrode, (b) conductivity, (c) bio affinity, and (d) active reaction sites.

conductivity or high bioaffinity [24, 25], (iii) increase in the three-dimensional structure of the electrode surface [26, 27], and (iv) nanometer metal or metal graft electrode surface [28, 29]. Refer to the surface modification method of anode electrode material [30]. These modification methods mainly improve the following three aspects (Figure 11.2): (i) conductivity, (ii) bioaffinity, and (iii) active reaction sites, in order to achieve the purpose of improving the synthesis efficiency of the microbial electrosynthesis system. The surface of the electrode is modified to have a positive charge, which can improve the interaction between the bacteria and the electrode, and is beneficial to the formation of the biofilm. Zhang et al. [29] used a variety of methods to modify carbon cloth and tested the performance of these electrode materials by Sporomusa ovata electrosynthesis. Compared with unmodified carbon cloth, chitosan and cyanuric acid modification increased the yield of acetic acid by six to seven times, and the modification of 3-aminopropyltriethoxysilane and polyaniline by three times. The treatment of metal nanoparticles (including gold, palladium, and nickel) increased the electrical synthesis efficiency by 6, 4.7, and 4.5 times, respectively. In addition, the electrode material obtained by modifying the cotton cloth and the polyester fiber with carbon nanotubes has an acetic acid electrosynthesis efficiency three times higher than that of the carbon cloth material. The yield of chitosan-modified electrode was the highest, reaching 45.8 ± 11.2 mmol/m2 /d, and the current density was 0.0475 mA/cm2 . The coulombic efficiency of all electrode materials is around 80%. Among these materials, the cost of metal nanoparticles (especially gold and palladium) and carbon nanotubes is relatively expensive. Electrochemical reduction of diazonium salt is a simple and stable reaction, which is modified by electrochemical reduction of specific functional groups to the surface of the electrode [31–33]. The electrode surface modification method has been successfully applied to the anode electrode material, which improves the current output of the microbial fuel cell (MFC) [31, 34]. Guo et al. [32] used electrochemical reduction of diazonium salt to graft –OH onto the surface of a glassy carbon (GC) electrode. The results show that the MFC of the –OH-grafted electrode has a significant increase in current. It is speculated that this is because

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the grafting of –OH increases the hydrophilicity of the electrode surface, the microorganisms are more likely to adhere to the electrode surface to form a tighter biofilm, and enrich more EAMs. Nie et al. [28] anchored nickel nanoparticles to the graphite surface and used the electrode material for microbial electrosynthesis of Sp. ovata. Compared with untreated graphite materials, the CO2 reduction rate of treated graphite materials was increased by 2.3 times and the acetic acid yield was 56.4 mmol/m2 /d. The nickel nanoparticle structure increases the cell attachment area and improves the conduction efficiency. In addition, positively charged nickel nanoparticles cause the negatively charged cells to adhere more closely to the surface of the graphite electrode, facilitating electron transport and acetic acid synthesis. Nanomaterials are the direction of electrode material development [35]. The reticular glass carbon has a microporous structure and is a good supporting material. Keller and coworkers [26] plated carbon nanotubes on the surface of reticulated glass carbon to produce a new multilayered hole electrode material NanoWeb-RVC (reticulated vitreous carbon). This new electrode material has good biocompatibility and conductivity, and its current density is three times higher than RVC. Using NanoWeb-RVC as the electrode material, the mixed culture was used for microbial electrosynthesis. The current density (normalized to projected area) reached 3.7 mA/cm2 , and the acetic acid yield (normalized to projected area) was 3.25 mol/m2 /d. Compared with the graphite plate electrode material, the ratio was increased by 1.7 and 2.6 times (normalized to the total area) Interestingly, the unmodified RVC surface has neither biofilm formation nor acetic acid formation, indicating that nanostructures are critical for biofilm formation and electron transport. The improvement in electrosynthesis performance is mainly due to the nanostructure of NanoWeb-RVC rather than the increase in total area, which increases bacterial–electrode interaction and extracellular electron transport rate. The high surface area/volume ratio feature of NanoWeb-RVC not only maximizes biofilm area but also increases mass transfer efficiency. NanoWeb-RVC also has great application potential in industrial-scale production. Firstly, the NanoWeb-RVC material is easy to scale and the cost is reasonable. Secondly, the support material RVC can be cut freely, so NanoWeb-RVC of different shapes and sizes can be prepared according to the needs of the reactor. The macroporous structure of RVC combined with the nanostructure of carbon nanotubes effectively increases the current transfer rate and the rate of microbial electrosynthesis. The unmodified RVC surface has neither biofilm formation nor acetic acid formation, indicating that nanostructures are critical for biofilm formation and electron transport. The improvement in electrosynthesis performance is mainly due to the nanostructure of NanoWeb-RVC rather than the increase in total area, which increases bacterial–electrode interaction and extracellular electron transport rate. The high surface area/volume ratio feature of NanoWeb-RVC not only maximizes biofilm area but also increases mass transfer efficiency. NanoWeb-RVC also has great application potential in industrial-scale production. The NanoWeb-RVC material is easy to scale and the cost is in a reasonable range. Secondly, the support material RVC can be cut freely, so different shapes can be prepared according to the needs of the reactor. The

11.3 Interspecies Electron Transfer Pathway

macroporous structure of RVC combined with the nanostructure of carbon nanotubes effectively increases the current transfer rate and the rate of microbial electrosynthesis. Keller and coworkers [36] electroplated a multiwalled carbon nanotube material on the surface of the reticulated glass carbon to prepare another novel electrode material named electrophoretic deposition (EPD)-3D, which is used for microbial electrosynthesis of mixed cultures. At a cathode potential of −0.85 V (vs. SHE), the current density reached 10.2 mA/cm2 and the acetic acid yield was 11.4 ± 0.5 mol/m2 /d. The effects of pH, electrode potential, and macropore diameter of electrode materials on electrosynthesis efficiency was explored [37]. An ultrahigh level of acetic acid yield of 22.2 mol/m2 /d–1 1330 g/m2 /d was achieved at a pH of 6.7, a pore size of 45 ppi (pores per inch), and a cathode potential of −1.1 V (vs. SHE). For an actual microbial electrosynthesis system, the choice of electrode materials is critical. The ideal electrode material needs to have a good electron active surface, a high electron transfer rate, and a low overpotential. In particular, for microbial electrosynthesis systems that directly deliver electrons, the electrode material also needs to have a large surface area and good biocompatibility to facilitate biofilm formation. In addition, the choice of electrode materials also requires cost and scalability. Conventional electrode materials are difficult to achieve efficient electron transfer between electrodes and microorganisms, and it is imperative to reform and develop new electrode materials.

11.3 Interspecies Electron Transfer Pathway IET refers to the phenomenon that the electron donor microorganism and the electron acceptor microorganism transfer electrons directly or indirectly to form a mutual growth relationship, thereby jointly completing a metabolic process that cannot be completed by a single microorganism (Figure 11.3) [38]. For example, in a coculture system of Escherichia coli with ethanol as an electron donor and fumaric acid as an electron acceptor, G. metallireducens cannot oxidize ethanol with fumaric acid as an electron acceptor, and G. sulfurreducens fumaric acid cannot be reduced with ethanol as an electron donor. However, when both microorganisms are present, G. metallireducens can transfer electrons produced by oxidation of ethanol to G. sulfurreducens and G. sulfurreducens and then transfer electrons to fumaric acid to realize electronic interaction between G. metallireducens and G. sulfurreducens [39]. Microbial IET is divided into mediated interspecies electron transfer (MIET) and direct interspecies electron transfer (DIET). MIET is a classic mutual growth mechanism, and the electron transfer between microorganisms is through redox endogenous (bacterial self-produced) small molecular substances such as hydrogen/formic acid, riboflavin, etc., or exogenous small molecular substances (naturally occurring or artificial synthesis) such as humus. DIET is a process of electronic exchange between microorganisms through their own conductive structures such as conductive pili and cytochrome c [40]. In addition, IET between microorganisms is mediated by exogenous conductive materials such as activated carbon

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11 External Electron Transfer: Pathway, Mechanism, and Microorganisms Involved

HCOOH H

H+

CO2

CO2

+

(b)

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e–



e

e–

e–

CO2

CO2

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(c)

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Organics e–

CO2

Organics

e–

e–

CO2

Fumarate

Succinate e–

e

Succinate

Organics e–

e

Organics –

ESred

Fumarate

Cell

Electrode

CH4

ESox –

Cell

H+ HCOOH

Cell

H+

Succinate

Organics

Electrode

H2

Electrode

(a)

H2

Cell

Electrode

CH4

Electrode

288

Cell CO2

Reduced sulfur e–

O2

H2O

e–

Cable bacteria filament

(f)

Figure 11.3 Mechanisms of microbial interspecies electron transfer. (a) Exchange of electrons between species via hydrogen/formate, (b) electron transfer shuttle, (c) electrically conductive pili, (d) proteins associated with outer cell surfaces, (e) conductive materials, and (f ) long-distance electron transport via redox coupling.

[41] and magnetite [42]. Such electron transport pathways do not require the assistance of an energy carrier but rely directly on the conductivity of the mediator for electron transfer, thus classifying such IETs as DIET. MIET includes interspecific hydrogen/formic acid transfer (i.e. the way in which IET is achieved by hydrogen/formic acid as an electron carrier) and electron shuttle mechanism (i.e. microorganisms use small molecular substances with redox properties such as phenazine, riboflavin, and humus). DIET includes nanowire mechanism (microbial use of self-conducting pili to achieve electronic cooperation), redox protein mechanism (mainly cytochrome c as electron carrier for electron transfer), and conductive substance-mediated mechanism (using exogenous conductive particles such as activated carbon, magnetite, etc., for electron transfer). In addition, long-range redox coupling between cable bacteria and sulfur-oxidizing bacteria found in marine sediments in recent years is also considered to be one of microbial IETs. Cable bacteria use their long filaments to transfer electrons from sulfur oxidizing bacteria to the oxygen-containing layer on the deposition surface, reducing oxygen to water, and electronically interacting with sulfur-oxidizing bacteria [43]. IET is ubiquitous and diverse in microorganisms. Electron transfer between Methanosaeta and Geobacter, which are widely found in methanogenic reactors and anaerobic sediments, can be achieved by means of hydrogen/formic acid [44] and redox small molecules [45]. They can also use the self-conducting pili and redox proteins for IET. At present, research on the use of hydrogen/formic acid for coculture between microorganisms is relatively mature, but DIET research is still in its infancy.

11.3 Interspecies Electron Transfer Pathway

11.3.1

Mechanism of MIET

In the interspecific hydrogen/formic acid transfer system, microbial oxidation of organic matter is usually associated with NAD+ /NADH, FAD/FADH2 or coenzyme F420/F420-H2 , and Fd(ox)/Fd(red) (Fd, ferredoxin). The Gibbs free energy of these coupling reactions in the standard state is usually positive, and the reaction does not occur spontaneously. When the partial pressure of hydrogen is less than a certain value, the Gibbs free energy changes from a positive value to a negative value, and the reaction can occur spontaneously, facilitating electrons from the electron donor microorganisms. The electron shuttle mechanism refers to the way in which electrons are transferred between microorganisms by means of self-secreting, naturally occurring or artificially synthesized small molecular substances (electron shuttles) having redox activity [46]. Biebl and Pfennig [47] found that sulfides can mediate microbial IET in coculture systems of Desulfuromonas acetoxidans and Prosthecochloris aestuarii. Among them, D. acetoxidans uses ethanol as an electron donor to reduce sulfur element to sulfide. P. aestuarii fixes CO2 by photosynthesis and oxidizes sulfide to element sulfur. In 2002, Kaden et al. [48] found l-cystine/cysteine-mediated IET in a coculture system of G. sulfurreducens and Wolinella succinogenes with acetic acid as electron donor and nitrate as electron acceptor. G. sulfurreducens reduces l-cystine to l-cysteine while oxidizing acetic acid, W. succinogenes uses l-cysteine to reduce nitrate, and l-cysteine is oxidized to l-cystine and is again involved in the oxidation of acetic acid. Smith et al. [49] found that the mutual metabolism rate of G. metallireducens and G. metallireducens was promoted in the case of exogenous addition of humic analog anthraquinone-2,6-disulphonic disodium salt (AQDS), and the coculture experiment of mutants proved that AQDS-assisted IET could be provide enough energy for G. sulfurreducens growth. 11.3.2

Mechanism of DIET

Summers et al. [39] found cytochrome c and pili-mediated DIET in a coculture system of G. metallireducens and G. sulfurreducens with ethanol as the electron donor and fumaric acid as the electron acceptor. The discovery of DIET broke the traditional perception of microbial IET and was a major milestone in microbial IET research. It is generally seen that the entangled structure formed by many pili is in the coculture of the DIET system in which the genus is involved, such as G. metallireducens and G. sulfurreducens. Under the action of pili, the mutual microbes form a close agglomerate to shorten the electron transfer distance and reduce energy loss. DIET is often not achieved for nonconductive pili-mediated mutual systems. For example, G. metallireducens cannot achieve IET with methanogen Methanosaeta or Methanosarcina in the absence of pili expression [50]. Moreover, none of the G. metallireducens and G. sulfurreducens knocked out the pili expression genes and could not interact with each other [39, 51], indicating the important position of pili in DIET. Pili is a filamentous protein that extends from the surface of cells and not only helps bacteria to attach but also promotes cell movement and transmit electrons to extracellular electron

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receptors. Rotaru et al. [50] used a modified atomic force microscope to find that G. sulfurreducens pili have a conductive ohmic response. Malvankar et al. [52] purified pili detection found that G. sulfurreducens pili conductivity can reach 188 mS/cm, which was like metal conductivity and proposed a metal-like conductivity model of pili. Sarah et al. [35] studied the process of G. sulfurreducens to transfer electrons to electrodes by techniques of gene knockout and proposed an electronic transition theory different from the metal-like conductivity model: electrons gradually transition to adjacent redox proteins along the pili. Electron acceptors achieve electron transfer, in which the pili plays a role as an electron transfer scaffold. Bonanni et al. [7] combined the metal-like conduction model and the electronic transition model to propose the Stepping stone theory: electrons are conducted along the aromatic ring clusters on the pili. When the aromatic ring cluster distance is greater than the direct transfer distance of electrons, the electrons are transmitted by the cytochrome c on the pili. 11.3.3

IET Microorganisms

The common forms of MIET are interspecies hydrogen/formic acid transfer and electron shuttle-mediated IET, in which interspecific hydrogen/formic acid transfer is common in methanogenic microbial communities, such as methanogens and acetogens, methanogens and Desulfovibrio, and Bacillus and methanol degrading bacteria. In addition, electronic shuttles such as sulfides, l-cysteine, AQDS, etc., can also assist microbial MIET, such as electron transfer between desulfurization bacteria and photosynthetic bacteria, bacillus and Wolinella. The electrogenic flora in mixed culture can be transported electronically by DIET (predation). From previous genetic and transcriptomic studies, G. metallireducens and G. sulfurreducens apparently metabolize ethanol to form conductive aggregates during syntrophism condition, further through the exchange of pili in the polymer [53]. The mechanism of DIET was also confirmed in anaerobic digester in the aggregates of G. metallireducens and Methanosaeta harundinacea. Granular activated carbon (GAC) is hypothesized to stimulate the feeding of bacteria and methanogens. GAC can affect the flagella of conductive bacteria and cytochrome c involved in the electrical connection between organisms in bacteria. Gene deletion studies indicate that DIET has GAC that does not require conductive fibers and associated c-type cytochromes [41]. The molecular mechanism of DIET and its contribution to energy production are not well understood, so further research is needed.

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(2011). Aryl diazonium salts: a new class of coupling agents for bonding polymers, biomacromolecules and nanoparticles to surfaces. Chem. Soc. Rev. 40 (7): 4143–4166. Picot, M., Lapinsonnière, L., Rothballer, M., and Barrière, F. (2011). Graphite anode surface modification with controlled reduction of specific aryl diazonium salts for improved microbial fuel cells power output. Biosens. Bioelectron. 28 (1): 181–188. Strycharz, S.M., Glaven, R.H., Coppi, M.V. et al. (2011). Gene expression and deletion analysis of mechanisms for electron transfer from electrodes to Geobacter sulfurreducens. Bioelectrochemistry 80 (2): 142–150. Jourdin, L., Grieger, T., Monetti, J. et al. (2015). High acetic acid production rate obtained by microbial electrosynthesis from carbon dioxide. Environ. Sci. Technol. 49 (22): 13566–13574. Jourdin, L., Freguia, S., Flexer, V., and Keller, J. (2016). Bringing high-rate, CO2 -based microbial electrosynthesis closer to practical implementation through improved electrode design and operating conditions. Environ. Sci. Technol. 50 (4): 1982–1989. Storck, T., Virdis, B., and Batstone, D.J. (2016). Modelling extracellular limitations for mediated versus direct interspecies electron transfer. ISME J. 10 (3): 621. Summers, Z.M., Fogarty, H.E., Leang, C. et al. (2010). Direct exchange of electrons within aggregates of an evolved syntrophic coculture of anaerobic bacteria. Science 330 (6009): 1413–1415. Stams, A.J. and Plugge, C.M. (2009). Electron transfer in syntrophic communities of anaerobic bacteria and archaea. Nat. Rev. Microbiol. 7 (8): 568. Liu, F., Rotaru, A.-E., Shrestha, P.M. et al. (2012). Promoting direct interspecies electron transfer with activated carbon. Energy Environ. Sci. 5 (10): 8982–8989. Cruz Viggi, C., Rossetti, S., Fazi, S. et al. (2014). Magnetite particles triggering a faster and more robust syntrophic pathway of methanogenic propionate degradation. Environ. Sci. Technol. 48 (13): 7536–7543. Vasquez-Cardenas, D., Van De Vossenberg, J., Polerecky, L. et al. (2015). Microbial carbon metabolism associated with electrogenic sulphur oxidation in coastal sediments. ISME J. 9 (9): 1966. Liu, Y., Xu, H.-L., Yang, S.-F., and Tay, J.-H. (2003). Mechanisms and models for anaerobic granulation in upflow anaerobic sludge blanket reactor. Water Res. 37 (3): 661–673. Von Canstein, H., Ogawa, J., Shimizu, S., and Lloyd, J.R. (2008). Secretion of flavins by Shewanella species and their role in extracellular electron transfer. Appl. Environ. Microbiol. 74 (3): 615–623. Roden, E.E., Kappler, A., Bauer, I. et al. (2010). Extracellular electron transfer through microbial reduction of solid-phase humic substances. Nat. Geosci. 3 (6): 417. Biebl, H. and Pfennig, N. (1978). Growth yields of green sulfur bacteria in mixed cultures with sulfur and sulfate reducing bacteria. Arch. Microbiol. 117 (1): 9–16.

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48 Kaden, J., Galushko, A.S., and Schink, B. (2002). Cysteine-mediated elec-

49 50

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52 53

tron transfer in syntrophic acetate oxidation by cocultures of Geobacter sulfurreducens and Wolinella succinogenes. Arch. Microbiol. 178 (1): 53–58. Smith, J.A., Nevin, K.P., and Lovley, D.R. (2015). Syntrophic growth via quinone-mediated interspecies electron transfer. Front. Microbiol. 6: 121. Rotaru, A.-E., Shrestha, P.M., Liu, F. et al. (2014). A new model for electron flow during anaerobic digestion: direct interspecies electron transfer to Methanosaeta for the reduction of carbon dioxide to methane. Energy Environ. Sci. 7 (1): 408–415. Shrestha, P.M., Rotaru, A.-E., Summers, Z.M. et al. (2013). Transcriptomic and genetic analysis of direct interspecies electron transfer. Appl. Environ. Microbiol. 79 (7): 2397–2404. Malvankar, N.S., Vargas, M., Nevin, K.P. et al. (2011). Tunable metallic-like conductivity in microbial nanowire networks. Nat. Nanotechnol. 6 (9): 573. Chen, S., Rotaru, A.-E., Liu, F. et al. (2014). Carbon cloth stimulates direct interspecies electron transfer in syntrophic co-cultures. Bioresour. Technol. 173 (Suppl C): 82–86.

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12 Extracellular Electron Transport of Electroactive Biofilm Xu Zhang Center for Microbial Ecology and Technology (CMET), Ghent University, Department of Biotechnology, Faculty of Bioscience Engineering, Coupure Links 653, 9000, Ghent, Belgium

12.1 Electroactive Bacteria Electroactive bacteria are able to exchange electrons with their external environment, which can include electrodes and minerals. Pure cultures and mixed populations of microorganisms have been demonstrated to grow on anodes or cathodes (Figure 12.1). Only a few microbes have been proven to directly transfer electrons from organics to a conductive surface. Bacteria capable of electron transfer with anodes include iron reducers (e.g. Geobacter sulfurreducens [2] and Shewanella oneidensis [3]), sulfur reducers (e.g. Desulfuromonas acetoxidans) [4], nitrate reducers (e.g. Comamonas denitrificans and Paracoccus pantotrophus) [6, 7], and phototrophic purple nonsulfur bacteria (e.g. Rhodopseudomonas palustris) [8], while with respect to cathodes, autotrophic microbes can transfer electrons from the electrode (or cathodically produced H2 ) to CO2 as the carbon source. Typical cathodic microbes include Sporomusa ovata, Sporomusa silvacetica, and Sporomusa sphaeroides in the class Negativicutes in the phylum Firmicutes, [9, 10] and Methanobacterium palustre [11] and Methanococcus maripaludis in the Euryarchaeota phylum of Archea [12]. 12.1.1 Role of Multiheme Cytochromes in Extracellular Electron Transport (EET) C-type cytochromes are major components of extracellular electron transfer process, of which the multiheme cytochrome pathway of S. oneidensis has been well studied. Electrons flow from a periplasmic decaheme cytochrome (MtrA) to another outer-membrane decaheme cytochrome (MtrC) through a transmembrane porin (MtrB) via the MtrCAB porin-cytochrome conduit [13]. The extracellular electron transfer of Shewanella might occur via functional cytochromes as “nanowires” or soluble electron shuttles [13]. Recently,

Bioelectrosynthesis: Principles and Technologies for Value-Added Products, First Edition. Edited by Aijie Wang, Wenzong Liu, Bo Zhang, and Weiwei Cai. © 2020 Wiley-VCH Verlag GmbH & Co. KGaA. Published 2020 by Wiley-VCH Verlag GmbH & Co. KGaA.

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12 Extracellular Electron Transport of Electroactive Biofilm

Electrons

Electrons

H2O CO2 O2 Organics N2

HS– S0

Cl–

Na+ K+ H+

U4+

NO3–

ClO4–

+

NH4

U6+

Anode

Cathode

Figure 12.1 Illustration of microbial anode and cathode, respectively. At the anode, the bacteria oxidize organic/inorganic compounds and deliver electrons to the electrode. At the cathode, electrons flow from the electrode to the bacteria, which reduces a chemical compound into more valuable ones (e.g. CO2 , O2 , and NO3 − ). Source: Rabaey et al. 2007 [1]. Reproduced with permission of Springer Nature.

nanowires of Shewanella were demonstrated to be composed of extensions of the outer membrane to enable long-distance electron transport [14]. Geobacter spp. possess abundant redox-active c-type cytochromes for the electron transferring across the cell membrane [15–18]. Unlike Shewanella, Geobacter does not release the soluble redox mediator to help in electron exchange with an insoluble electron acceptor/donor [19]. Geobacter can form electrical conductive biofilm matrix, and the conductivity of the biofilm can be considered as a combination of bacterial nanowires, polysaccharides [20], and several c-type cytochromes [15, 18, 20–24]. These bacterial nanowires consist of a repeating arrangement of pilA protein, which are anchored to the cell envelope, although their nature and diversity are still under discussion [25]. All these elements have been implicated in the extracellular electron transport (EET) between microorganisms and conductive surface, but their explicit roles in the process remain unclear. Electrons not only need to be transferred outside the cell but also transported over a long-distance conductive matrix for a distant (up to tens of micrometer) terminal electron acceptor (i.e. further than the typical length of a single cell). Different mechanisms have been proposed: “metallic-like” conductivity along microbial nanowires, “redox conduction” by electron self-exchanging between immobilized redox cofactors, and soluble redox mediators (Figure 12.2). For the later part, the discussion will mainly consider Geobacter or Geobacter-dominated anodic electroactive biofilms (EABs) as the example.

Medred Medox

DET

Diffusion

Redox conduction

Metallic-like conduction

Biofilm

12.2 Electron Transport Across Geobacter(−Dominated) EABs

Electron acceptor Cytochrome

Conductive pilus Medred/ox Molecular mediator

Figure 12.2 Schematic illustration of different proposed extracellular electron transfer mechanisms through an anodic biofilm. The so-called “direct electron transfer” (DET) along the redox enzymes anchored to the cell membrane that is in contact with the conductive surface, “metallic-like” conductivity via conductive pili, redox conduction via electron hopping between extracellular redox proteins, and mediated electron transfer based on endogenous or exogenous dissolved redox molecules. Source: Schröder and Harnisch 2017 [26]. Reproduced with permission of Elsevier.

12.2 Electron Transport Across Geobacter(−Dominated) EABs The mechanisms of electron transfer from electrode to the microbes are unclear, while the microbes to electrode mechanisms are more understood. Electrons generating from intracellular metabolism are transferred from the inner cell across the cell membrane and then transported along the extracellular conductive matrix (e.g. via proposed metallic-like conductivity or redox conduction) until they reach the conductive surface [27, 28]. Because classic EABs can form relatively thick biofilms (up to 100 μm vs. 1 μm for a single cell length), a better understanding about the long-distance EET is necessary. The aforementioned two mechanisms, “metallic-like” conductivity and redox conduction, have been considered as primary ones. It must first be noticed that the different hypotheses for conduction mechanisms along EABs are not necessarily exclusive, and multiple combinations of conduction couple be in parallel. 12.2.1

“Metallic-like” Conductivity via Microbial Nanowires

Similar to polyaniline, the overlapping π-orbitals of aromatic residues in type IV pili have been illustrated as “metallic-like” conductivity to living Geobacter biofilms [25, 29]. The conductivity of the individual isolated pili increases as temperature decreases, as a metallic-like temperature dependence [29, 30]. Early related studies only characterize the pilus conduction mechanism under certain conditions (vacuum and dry conditions) instead of physiologically relevant

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conditions. Recently, a linear relationship between current and voltage of purified pili films in hydrated conditions were demonstrated as ohmic conduction [31], while for live Geobacter biofilm characterization, the electrons could only transport across living biofilm when the interdigitated electrode potentials were close to apparent mid-point potential of microbial redox cofactor. The latter result supports a previous study suggesting that ohmic conduction alone could not percolate a hydrated, living G. sulfurreducens biofilm and that an additional redox conduction must be involved [32]. Recently, the electrical conductive filament of G. sulfurreducens has been demonstrated to be composed of polymerized multiheme cytochromes (OmcS) [33]. 12.2.2

Redox Conduction

The extracellular multiheme cytochromes in EABs can be considered as the essential part of redox cofactors regarding the redox polymers [34]. The distance between neighboring redox sites has been assumed close enough for a sequence of self-exchange reactions. A rectangular-box-shaped EAB was used as a simplified schematic representation to illustrate the proposed model (Figures 12.3 Direction of electron flux A

B

Δ

C

E

D

Z

Y X

e– X–

X–

X–

δ 1

2

3

Figure 12.3 Illustration of redox conduction via electron hopping across an idealized three-dimensional cubic lattice of immobilized redox sites (reduced sites, black circles; oxidized sites, white circles). Insets illustrate a counter-ion (X− ) associated with a single electron hopping event between neighboring redox sites. Δ denotes the cubic lattice parameter, whereas 𝛿 denotes electron hopping distance (𝛿 = Δ if the redox sites are fully immobilized; 𝛿 < Δ if Brownian motion occurs). Source: Dalton et al. 1990 [35]. Reproduced with permission of Elsevier.

12.2 Electron Transport Across Geobacter(−Dominated) EABs

jcat

Anode

Cell Cred Cox

jcat (a) Anode

8e–

CH3COOH + 2H2O 2CO2 + 8H+ j Cell Cred Cox

j (b)

0

d

Figure 12.4 Simplified illustration of long-distance EET through a multicell thick conductive biofilm matrix as the redox conduction under (a) turnover condition (with acetate as electron donor) generating a quasi-steady state catalytic current jcat and (b) nonturnover condition (acetate depletion) generating transient currents related to the charge and discharge of redox cofactors within the EAB. Source: Image modified from Boyd et al. [34].

and 12.4) [34]. Here, eight electrons were generated when electroactive bacteria oxidized one acetate molecule. These electrons are then delivered across the cell membrane to the surrounding extracellular matrix. An oxidized redox cofactor can accept an electron(s) and become (fully or partially) reduced and vice versa (C ox + e ⇌ C red ). The electron continues to be delivered from reduced to oxidized cofactors along the conductive biofilm matrix. The local flux of electrons is driven by the concentration gradient of redox states of redox cofactors as a diffusive process. Once the electron reaches the interface, the heterogeneous electron transfer can be considered as quasi-reversible (e.g. nonlimiting the current generation [32, 36]). While a definitive molecular mechanism(s) of electron transport across EABs is (are) not fully understood and still debated, abundant electrochemical data under diverse methodologies are well described by models based on redox conduction. The peak-shaped dependency of source-drain current response on the electrochemical gating measurements strongly suggests that distant EET across G. sulfurreducens biofilms grown on gold interdigitated microelectrode array (IDA) was driven by a redox gradient [32]. This result suggests that a redox-mediated process is necessary for long-distance conductance within the EAB (Figure 12.4), while not excluding a local impact of ohmic conduction on a smaller scale. In another study, an Arrhenius relationship between the electrical conductivity and temperature of living Geobacter biofilm has been reported, and the corresponding activation energy (0.13 eV) is close to those recorded

299

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for electron self-exchange between c-type cytochromes [37]. Comparing live Geobacter biofilm to the different conductive materials, Phan et al. further demonstrated that the electrical properties of live Geobacter biofilms are similar to those of hydrated redox polymers than those of ohmic conductors [38]. Besides the two primary mechanisms, another mixed mechanism, “stepping stones,” has been proposed in which the outer membrane cytochromes bonded to pili transfer the electron over a long distance [39]. 12.2.3

Basic Electrochemical Characterization of Redox Conductors

Cyclic voltammetry (CV), electrochemical impedance spectroscopy (EIS), double-potential step chronoamperometry (DPSC), and open-circuit potential (OCP) techniques are among the conventional methods to characterize the redox state and the electron transport across redox polymers. “Semi-infinite diffusion” is an analogous model for redox polymer films. It can be used to assess the long-range electron transport across EAB, while considering some experimental criteria, i.e. performing sufficiently fast electrochemical measurements with sufficiently thick biofilms. If the migration effect can be eliminated from electrons transport process, the evolution of the current density over time, j(t) , can then be derived from a one-dimensional Fick’s first law (along the distance normal to the electrode surface, x) for the electron flux at the interface between electrode and EAB (x = 0): ( ) dCred j(t) = F × J(t, x=0) = F × Dapp dx t, x=0 where J (t) is the flux of electrons over time across the electrode/biofilm interface (mol/cm s), F is the Faraday constant (96 485 C/mol), Dapp is the apparent diffusion coefficient of electron (cm2 /s), and C red is the concentration of reduced redox cofactors in the EAB (mol/cm3 ) (we assume one electron per redox cofactor to simplify the model). CVs are often used to characterize the electron transfer interaction at the interface between electroactive microorganisms and conductive surface. In a classic CV, the potential of the electrode starts from an initial potential (Ei ), then sweeps linearly to a final potential (E𝜆 ). During the interval, an electrochemical reaction is illustrated as Faradic current (e.g. a transient peak if mass transport is limiting and bulk concentration of electroactive species decreasing). The potential then sweeps back to Ei , accompanying with an opposite direction of electron transfer and reformation of initial redox states of the cofactors. An apparent midpoint potential (E1/2 ) can be extracted from the cyclic voltammogram. Figure 12.5 shows the principle of typical CV recorded under nonturnover conditions (absence of substrate) of a Geobacter-dominated EAB. If we assume the electron transport as a semi-infinite diffusional process, along a sufficiently thick EAB under nonturnover conditions (substrate depletion), the peak current density is proportional to the square root of potential scan rate(v), described by the Randles–Ševˇcík equation: ) ( nFvDapp 1∕2 jp = 0.446nFCT RT

12.2 Electron Transport Across Geobacter(−Dominated) EABs

180

E/V Eλ

Cred – e → Cox jpa

j μA/cm

90

0

Eλ Ei

–90 Ei

E1/2 jpc Cred → Cox + e

0

Switching time

(a)

–180 –0.8

t (b)

40 mV/s –0.3 E / V vs. Ag/AgCl

0.2

Figure 12.5 (a) Principle of a cyclic voltammetry. (b) Recorded data with a mature EAB under non-turnover conditions on a glassy carbon electrode; scan rate 40 mV/s. Source: Zhang et al. 2017 [36]. Reproduced with permission of John Wiley & Sons.

Here, we can extract the apparent electron diffusion coefficient (Dapp ) or the product (C T × D1/2 app ) to characterize the diffusional electron transport process. The value of the charge transport product C T × D1/2 app can be extracted from the slope of f (jp ) = f (v1/2 ) when the concentration of redox cofactors C T is not accessible. Besides CV, another traditional electrochemical analysis, DPSC, has often been applied to characterize and study the electron diffusion performance of abiotic redox polymers. Here, we use it to assess EET along the EAB. If the average concentration of redox cofactors throughout the biofilm is constant, and the heterogeneous electron transfer is (quasi-)reversible, the electrochemical response resembles that of semi-infinite diffusion, assuming that the redox-mediated electron transport is purely driven by diffusion. As demonstrated in Figure 12.6, before t = 0, the potential of the electrode is poised at Ei , a sufficiently low potential to fully reduce cofactors in the region close to the electrode surface. Once the potential has changed to Ef (after t = 0), only the reduced redox cofactors are available to be oxidized at the electrode. Hence, a large quantity of electrons flow toward the electrode, which reflects the electron depletion along the biofilm matrix. A linear relationship between the Faradic current response and t −1/2 is predicted, which is known as the Cottrell relation (including anodic and cathodic process): j(t) = ±

nFACT D1∕2 (πt)1∕2

The Cottrell relation can be applied in the opposite direction. Before t = 0, it poised the potential at a sufficient positive value to fully oxidize the redox cofactors close to the interface. Then stepped to a potential that is low enough to fully reduce the oxidized form of redox cofactors. With the current response, the charge transport product C T × D1/2 app can also be extracted in the reversible process.

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12 Extracellular Electron Transport of Electroactive Biofilm

E/V Ef

(–)

302

E1/2 Ei t

(a) j

j(t) =

nFACTD1/2 (π t)1/2

t

(b)

Figure 12.6 Principle of double-potential step chronoamperometry for an electron discharge process. (a) Initial step at negative potential E i to “fully” reduce the redox cofactors of the EABs (e.g. at least 120 mV lower than the midpoint potential, E 1/2 ). The second step is the reoxidation of the redox cofactors at a potential E f allowing a full reoxidation (e.g. at least 120 mV higher than E 1/2 ). (b) Illustration of a typical current response: the current decreases over time as the effect of depleting the built-up redox-active species near the surface. The Faradic current during the final step is assumed as a purely diffusive process followed Cottrell equation.

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conductor: a systematic study of the moisture and temperature dependence of its electrical properties. Phys. Chem. Chem. Phys. 18 (27): 17815–17821. 39 Bonanni, P.S., Massazza, D., and Busalmen, J.P. (2013). Stepping stones in the electron transport from cells to electrodes in Geobacter sulfurreducens biofilms. Phys. Chem. Chem. Phys. 15 (25): 10300–10306.

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13 Microbial Growth and Ecological and Metabolic Characteristics in Bioelectrosynthesis Systems Qian Liu and Sihao Lv Dongguan University of Technology, Research Center for Eco-Environmental Engineering, No.1, Daxue Road. Songshan Lake, Dongguan, Guangdong Province 523808, China

Microbial electrosynthesis (MES) is an emerging technology platform to directly produce carbon-neutral fuels and biochemicals from electricity, CO2 , and even waste organics [1, 2]. In MES systems, exoelectrogenic microorganisms are important catalysts for oxidation and reduction reactions that occurred at electrodes [3, 4], affecting product spectrum (such as alcohols, methane, organic acids, and so on) [5–8] and transformation efficiency.

13.1 Microbial Growth Kinetics and Energetics 13.1.1

Stoichiometry of Microbial Growth Systems

In microbial electrosynthesis (MES), three processes were performed by microbial community, including microorganism’s growth and biomass formation, maintenance of cellular function, and production of desirable biochemical (Figure 13.1). Stoichiometric analysis can be used to assess energy (adenosine triphosphate [ATP], H2 , or electron) requirements of microorganisms and their capability of producing biochemicals or biofuels, which will provide basis for engineered bacteria by using synthetic microbiology [9]. Although microbial reactions (e.g. growth processes and product formation) obey the laws of thermodynamics, the complexity of life activity limited the thermodynamic application of bioprocesses [10]. Gibbs free energy is considered a driving force for microorganism growth and metabolism, which is accompanied by enthalpy and entropy variation (as is shown in Eq. (13.1)). G = ΔH − TΔS

(13.1)

Here, G, H, S, and T represent Gibbs free energy, enthalpy, entropy, and temperature, respectively.

Bioelectrosynthesis: Principles and Technologies for Value-Added Products, First Edition. Edited by Aijie Wang, Wenzong Liu, Bo Zhang, and Weiwei Cai. © 2020 Wiley-VCH Verlag GmbH & Co. KGaA. Published 2020 by Wiley-VCH Verlag GmbH & Co. KGaA.

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13 Microbial Growth and Ecological and Metabolic Characteristics in Bioelectrosynthesis Systems

Figure 13.1 The main processes during microbial growth.

Catabolism Electron donor Electron acceptor Oxidized donor Reduced acceptor Energy

Carbons source Nitrogen source

Maintenance

Biomass

Anabolism

Microbial metabolic process contained catabolism and biosynthesis, which are reflected by Gibbs free energy dissipation (Eq. (13.2)). 0′

ΔR G =

n ∑





0 0 𝜈i,Products Δf Gi,Products + 𝜈i,Educts Δf Gi,Educts

(13.2)

i=1 ′

ΔR G0 is the Gibbs free energy of reaction for biochemical standard conditions; ′ Δf G0 is the Gibbs free energy of reactants formation in biochemical standard conditions (1 mol/l of respective reactants, 298.15 K, 101.325 kPa, pH = 7), ′ ′ kJ/mol. Standard Gibbs free energy of Δf G0 and standard enthalpy Δf H 0 for potential metabolic compounds in MES are listed in Table 13.1 [11]. However, Gibbs free energy should be corrected for nonstandard temperature or concentrations conditions in actual operation. Recent studies have shown that temperature is another key factor influencing extracellular electron transfer (EET) [12, 13]. Temperature correction and concentration correction are performed by Eqs. (13.3) and (13.4) [11]. T −T ′ ′ ′ T + ΔR H 0 (T) S ΔR G0 (T) = ΔR G0 (TS ) T T ) S ( S∑n ′ i=1 Ci,Products ΔR G = ΔR G0 + RT ln ∑n i=1 Ci,Educts

(13.3) (13.4)

13.1 Microbial Growth Kinetics and Energetics

Table 13.1 Standard Gibbs energy and enthalpy of compounds associated in MES. ′

Name

Structure

Phase

𝚫f G 0 (kJ/mol)

Bicarbonate

HCO3−

aq

−586.9



𝚫f H 0 (kJ/mol)

−692.0

Carbon monoxide

CO

g

−137.2

−110.5

Carbon dioxide

CO2

g

−394.4

−393.5

Ammonium

NH4+

aq

−79.4

−133.3

Proton

H+

aq

−39.9

0

Hydrogen

H2

g

0

0

Methane

CH4

g

−50.8

−74.8

Formate

CHO2−

aq

−335.0

−410.0

Acetate

C2 H3 O2−

aq

−369.4

−485.8

Ethanol

C2 H5 O

aq

−181.8

−288.3

Glycerol

C3 H8 O3

aq

−488.5

−676.0

Butyrate

C4 H7 O2−

aq

−352.6

−535.0

Glucose

C6 H12 O6

aq

−919.8

−1264.2

Succinate

C4 H4 O4 2−

aq

−690.2

−909.0

Lactate

aq

−517.1

−687.0

Tartrate

C3 H5 O3− C4 H4 O6 2−

aq

−1010.0



Biomassa)

CH1.8 O0.5 N0.2

s

−67.0

−91.0

a) The structure of biomass can be represented by CH1.8 O0.5 N0.2 because most microorganisms have the similar relative contents of 40–70% protein, 1.2% DNA, 5–15% RNA, 2–10% lipid, and 3–10% carbohydrate.

In the above equations, T is the actual temperature, T S is the standard tem′ perature, ΔR H 0 is the standard enthalpy, and C i is the concentration of the ith reactant. Electron transfer is considered a severe energetic loss for electroactive microorganisms and microbial electrochemical systems. It is known that polymerization, carboxylation, and Pi-addition reactions required energy input. Catabolism is the process of providing energy in which an electron acceptor (EA) and electron donor (ED) undergoes redox reaction, storing energy in the form of ATP to drive anabolism. The derived energy of Geobacter anodes was 27 ± 6 kJ/mol of electrons, which is in the order of the energy generated from hydrolysis of 1 mol ATP to adenosine diphosphate (ADP) [13]. In the reactions of mediated electron ′ transfer (MET), ΔR G0 of redox mediators can be calculated by standard potentials (Eq. (13.5)), and the difference of microbial growth stoichiometry was mostly determined by the electron donor and acceptor couples used [11]. The standard ′ potential is referred to by E0 , the number of electron transfer is represented by z, and Faraday constant is F. ′



ΔR G0 = −E0 zF

(13.5)

However, for MES with direct electron transfer (DET), EA received electron from cathode so that no ED couple was assigned straightforwardly. Little study

311

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13 Microbial Growth and Ecological and Metabolic Characteristics in Bioelectrosynthesis Systems

was focused on formal potential of EET of biocathodes [14]. In that case, the stoichiometry of microbial growth systems was different from MET. Research studies have demonstrated that the mechanisms of cathodic EET were different from bioanodes, and the energy generated per transferred electron was lower than bioanodes [15]. 13.1.2

Electrode-Respiring Bacteria Kinetics

Bioelectrosynthesis system is a system of mixing various organic substances. There are a variety of EET mechanisms between the microorganism and the electrode, including direct contact through the outer membrane (OM) protein, diffusion of the soluble electron shuttle, and electron transport of the extracellular biofilm matrix through the solid. Therefore, the kinetic parameters include anode overpotential, maximum current density, and Monod’s half-saturation constant. Torres et al. [16] measured the current density at different anode potential substrates at different concentrations of a single substrate using continuous flow dual-chamber h-type microbial electrolysis cells (MECs). The results showed that butyrate and hydrogen produced lower current densities, and propionate, acetate, and ethanol produced higher current densities (a maximum of 1.6, 9.0, and 8.2 A/m2 , respectively). Acetate produced a high coulombic efficiency (CE) of 86%, whereas ethanol and propionate had a CE of 49% and 41%, respectively. Anode-respiring bacteria (ARB) undergo a respiratory metabolic process on the anode in an electrosynthesis system. Because in an electrosynthesis system the electrode is a solid conductor, it can neither dissolve nor reduce but only acts as a conductor. ARB bacteria can only use the EET mechanism to obtain additional electrons for respiration. To date, ARB include members from diverse phyla, such as Alpha-, Beta-, Gamma-, and Deltaproteobacteria, Firmicutes, Acidobacteria, and yeast [17]. Because ARB bacteria are at a certain distance from the anode surface, the total anode potential loss (𝜂 anode ) is defined as the difference between the electron-donor potential and the anode potential (𝜂 anode = Edonor − Eanode ) as shown in Figure 13.2. The energy loss associated with intracellular potential determines the energy (potential, E) of electrons released from OM cells through the OM cytochrome into the EET mechanism (EOM ). The loss of ARB potential is divided into endogenous loss and exogenous loss, where endogenous loss includes substrate utilization and electron generation, Edonor

EET

Eanode

Cell Cell Cell

EOM

Edonor

e–donor

Intracellular potential losses Cell

E CO2 + H+

Einterface

Figure 13.2 Electroactive biofilm potential loss.

Extracellular potential losses

EOM Einterface

Eanode

Distance from the anode

13.1 Microbial Growth Kinetics and Energetics

while exogenous loss includes electron transfer between bacteria and electrodes, and the loss of electrons transmits on solid surface, and the surface electrons are transmitted to the electrodes. The Monod equation is generally used to describe the microbial substrate utilization process. The Monod equation can accurately describe the relationship between bacterial oxidative substrate rate and reducing intracellular electron carriers, it is generally used to evaluate the microbial substrate utilization process. The current density produced by the ARB in the biofilm can be expressed by the Eq. (13.6) [18]. j = jmax

S

(13.6)

ks,app + S

where j is the current density obtained by ARB, jmax is the maximum current density of the ARB biofilm, S is the substrate concentration in the liquid, and K s,app is the apparent half-saturation substrate concentration in a biofilm. When the electron donor and acceptor are both soluble, the multiplicative Monod model equation (13.7) can also be used to express the substance utilization. Sd Sa (13.7) q = qmax 𝜑a Sd + KSd Sa + KSa q is the specific rate of ED utilization (mmol-ED mg/VS d; VS is volatile solids, a measure of biomass), Sd is the ED concentration (mmol-ED/cm3 ), KSd is the half-max-rate ED concentration (mmol-ED/cm3 ), Sa is the EA concentration (mmol-EA/cm3 ), KSa is the half-max-rate EA concentration (mmol-EA/cm3 ), 𝜙a is the volumetric fraction of active biomass (dimensionless), and qmax is the maximum specific rate of ED utilization (mmol-ED mg/VS d). ARB bacterial electron production rate often uses Nernst–Monod Eq. (13.8) [18]: ⎛ ⎞ ⎜ ⎟ 1 j = jmax ⎜ [ ]⎟ F ⎜ 1 + exp − (EOM − EKA ) ⎟ RT ⎝ ⎠

(13.8)

where R is the ideal gas constant (8.3145 J/mol K), F is the Faraday constant (96 485 C/mol e), T is the temperature (K), and EKA is the potential at which j = 1/2jmax (V). Because the Nernst–Monod equation is derived from the Monod equation, it is able to describe the irreversible metabolic response of microorganisms under steady-state conditions. Different kinetic equations can be used to describe the two potential loss pathways of the bacterial exogenous electron transport process. Electrons are transported from the bacterial outer membrane protein to the electrode surface by the electronic shuttle, or by a solid conductive matrix. Therefore, the Fick’s law, Ohm’s law, and Butler–Volmer equations can be used to describe these processes. Fick’s law j = nF

(

Dshuttle ΔCshuttle Δz

) (13.9)

313

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13 Microbial Growth and Ecological and Metabolic Characteristics in Bioelectrosynthesis Systems

where Dshuttle is the diffusion coefficient of the electron shuttle (m2 /s), Δz is the transport distance (m), ΔC shuttle is the concentration gradient of either oxidized or reduced shuttle (mol/m3 ), and nF converts from moles to coulombs. Ohm’s law Kbio (EOM − Einterface ) (13.10) Δz where K bio is the conductivity of the solid conductive matrix (S/m). K bio determines the potential losses (EOM − Einterface ) associated with j. j=−

Butler–Volmer j = −j0 exp

[

nF(1 − 𝛼)(Eanode − E0 ) RT

] (13.11)

where j0 is the exchange current density (A/m2 ), 𝛼 is the electron transfer coefficient or the symmetry coefficient for the anodic or the cathodic reaction, 0 Eanode is the anode potential (V), and Einterface is the standard potential (V) of the reaction occurring at the anode interface. Because the electron transfer process is subdivided, the biofilm conduction model could be built to evaluate the kinetics of the electron transfer mechanism. Lee et al. [19] assessed the kinetics of electron transfer from acetate to the anode for a mixed culture biofilm anode. The models consist of three parts: (i) intracellular electron transfer, (ii) non-Ohmic EET from an outer membrane protein to an extracellular cofactor, and (iii) electron transfer from the extracellular cofactor to the anode by Ohmic conduction in the biofilm matrix. Lee set up a miniature microbial electrochemical cell whose steady-state current density was 0.82 ± 0.03 A/m2 at a fixed anode potential of −0.15 V vs. the standard hydrogen electrode (SHE). If Ohmic conduction EET was limited, the maximum current density could be as high as 270 A/m2 when Geobacter genus was