Integrated Microbial Fuel Cells for Wastewater Treatment 0128174935, 9780128174937

Table of contents :
Cover
Integrated Microbial Fuel Cells for Wastewater Treatment
Copyright
Contents
List of contributors
Part 1: Introduction
1 Introduction to microbial fuel cells: challenges and opportunities
Chapter Outline
1.1 Introduction
1.2 Brief history of microbial fuel cells to bioelectrochemical systems
1.3 Principles and challenges of microbial fuel cells
1.4 Future of microbial fuel cells
1.5 Conclusion
Acknowledgment
References
2 Microbial fuel cell–integrated wastewater treatment systems
Chapter Outline
2.1 Introduction
2.1.1 Sediment microbial fuel cells
2.1.2 Constructed wetlands-microbial fuel cells
2.1.3 MBR-microbial fuel cells
2.1.4 Desalination cell-microbial fuel cells
2.1.5 Other processes
2.2 Conclusion
References
Part 2: Application to Industrial Wastewater treatment
3 Removal of heavy metals using bioelectrochemical systems
Chapter Outline
3.1 Introduction
3.2 Bioelectrochemical systems for heavy metal removal
3.2.1 Concept and principle
3.2.2 Reduction of heavy metals at the cathode of bioelectrochemical systems
3.3 Electrode materials used for heavy metal removal in bioelectrochemical systems
3.4 Conventional technologies versus bioelectrochemical systems-based technology for the removal of heavy metals
3.5 Conclusion
References
Further reading
4 Textile wastewater treatment using microbial fuel cell and coupled technology: a green approach for detoxification and bi...
Chapter Outline
4.1 Microbial fuel cell and its application in the treatment
4.1.1 Mechanisms involved in dye breakdown
4.1.2 Dye removal and current generation in microbial fuel cell
4.1.3 Dye removal and total COD removal
4.2 Enhancement of microbial fuel cell performance
4.2.1 Bioanode-based enhancement of dye treatment
4.2.2 Biocathode-based enhancement of dye treatment
4.2.3 Membrane-based enhancement of dye treatment
4.2.4 Effect of the shuttle on dye removal and electricity generation
4.3 Microbial diversity involved in the breakdown of dye in microbial fuel cell
4.4 Toxicity of treated dye wastewater
4.5 Microbial fuel cell–coupled techniques for textile wastewater treatment
4.5.1 Microbial fuel cell–integrated constructed wetlands
4.5.2 Microbial fuel cell couple aerobic biocontact oxidation reactor system
4.5.3 Bioelectro-Fenton technology-microbial fuel cell
4.5.4 Electrolysis cell combined with a microbial fuel cell (MFC-MEC)
4.6 Research gap
Acknowledgments
References
Further reading
5 Agro-industrial wastewater treatment in microbial fuel cells
Chapter Outline
5.1 Introduction
5.2 Use of agro-industrial wastewater as substrate for microbial fuel cells
5.3 Dairy industry wastewater
5.4 Brewery and winery industry
5.4.1 Brewery wastewater
5.4.2 Winery wastewater
5.5 Agro-industrial wastewaters and by-products
5.5.1 Palm oil industry wastewater
5.5.2 Agricultural products processing wastewater
5.5.3 Agricultural residues
5.6 Livestock industry wastewater
5.7 Challenges in using microbial fuel cells
5.8 Conclusion
References
6 Pharmaceutical wastewater treatment in microbial fuel cell
Chapter Outline
6.1 Introduction
6.2 Application to pharmaceutical wastewater treatment
6.3 Integration of microbial fuel cell with other wastewater-treatment processes
6.4 Large-scale microbial fuel cell: potentials and challenges
References
7 Oil and petrochemical industries wastewater treatment in bioelectrochemical systems
Chapter Outline
7.1 Introduction
7.2 Oil field and petrochemical wastewater treatment in the conventional treatment process
7.3 Oil field and petrochemical wastewater treatment in the bioelectrochemical system
7.4 Conclusion
References
Further reading
8 Bioelectrochemical systems for stormwater treatment and energy valorization processes
Chapter Outline
8.1 Introduction
8.1.1 Urban stormwater
8.2 Energy recovery and stormwater treatment efficiency with bioelectrochemical systems
8.2.1 Influencing factors for bioelectrochemical system
8.2.1.1 pH
8.2.1.2 Temperature
8.2.1.3 Electroconductivity
8.2.1.4 Kinetics and thermodynamics
8.2.1.5 Electron transfer mechanism
8.2.1.6 Electrodes and applied potential
8.2.1.7 Membranes
8.3 Economic and environmental considerations
8.3.1 Environmental economics
8.3.2 Environmental impact and life cycle analysis
8.4 Technical scales of bioelectrochemical systems
8.5 Outlook, challenges, and future perspectives
8.6 Conclusion
References
Further reading
9 Treatment of food processing and beverage industry wastewaters in microbial fuel cells
Chapter Outline
9.1 Introduction
9.2 Food-based wastes and wastewater as a substrate for microbial fuel cell
9.3 Beer brewery wastewater wastes and wastewater as a substrate for microbial fuel cell
9.4 Conclusion
References
Further reading
Part 3: Integration of MFC with other wastewater treatment processes
10 Microbial fuel cell coupled with microalgae cultivation for wastewater treatment and energy recovery
Chapter Outline
10.1 Introduction
10.1.1 Microbial fuel cells
10.1.2 Microalgae cultivation
10.2 Microbial fuel cell and microalgae cultivation–based integrated systems
10.2.1 Microbial fuel cells coupled with the algal photobioreactors
10.2.2 Microbial fuel cells with the algal biocathodes
10.3 Factors influencing the performance of integrated microbial fuel cell and microalgae cultivation systems
10.3.1 Light intensity
10.3.2 Carbon dioxide
10.3.3 pH
10.3.4 Dissolved oxygen
10.4 Conclusion
Acknowledgment
References
11 Integration of bioelectrochemical systems with other existing wastewater treatment processes
Chapter Outline
11.1 Introduction
11.2 Integration of bioelectrochemical system with electro-Fenton process
11.3 Integration of bioelectrochemical system with aerobic processes
11.4 Integration of microbial fuel cell with anaerobic digestion
11.5 Microbial fuel cell integration with septic tank
11.6 Microbial fuel cell integration with dark fermentation
11.7 Microbial fuel cell integration with microalgae
11.8 Novel integration of other processes with microbial fuel cells
11.9 Way forward
References
12 An overview of membrane bioreactor coupled bioelectrochemical systems
Chapter Outline
12.1 Introduction
12.1.1 Wastewater and its sources
12.1.2 Conventional wastewater treatment practices and lacunas
12.2 Bioelectrochemical systems
12.2.1 Evolution tree of bioelectrochemical systems
12.2.2 Major forms of bioelectrochemical systems
12.3 Membrane bioreactor
12.4 Hybrid bioelectrochemical system–membrane bioreactor systems: principle, treatment efficiency, and performance index
12.4.1 Integrated bioelectrochemical system–membrane bioreactor systems
12.4.1.1 The membrane as cathode-cum-filtration unit
12.4.1.2 The membrane as anode-cum-filtration unit
12.4.1.3 The membrane as separator-cum-filtration unit
12.4.2 Combined bioelectrochemical system–membrane bioreactor system
12.4.2.1 Membrane bioreactor as pretreatment unit
12.4.2.2 Membrane bioreactor as post-treatment unit
12.5 Outlook and future perspectives
12.5.1 Water-energy nexus
12.5.2 Membrane fouling mitigation
12.5.3 Control of emerging contaminants
12.5.4 Field-scale applications
12.6 Conclusion
Acknowledgment
References
13 Integration of microbial fuel cell into constructed wetlands: effects, applications, and future outlook
Chapter Outline
13.1 Introduction
13.1.1 Probable electron transfer mechanism in constructed wetlands-microbial fuel cell
13.1.2 Basic characteristic of constructed wetlands and their similarity with microbial fuel cell
13.2 Development of merger technology
13.2.1 Design and operation of constructed wetland-microbial fuel cells
13.2.2 Performance assessment of constructed wetland-microbial fuel cells
13.3 Challenges and future perspectives
Acknowledgment
References
14 Microbial fuel cell coupled with anaerobic treatment processes for wastewater treatment
Chapter Outline
14.1 Introduction
14.2 Integration of microbial fuel cell in anaerobic digestion
14.3 Microbial fuel cell coupling to treat undigested organics in the effluents of anaerobic digestion
14.4 Microbial fuel cells coupled anaerobic digestion for nutrient recovery and toxicity removal
14.5 Microbial fuel cell coupling in anaerobic digestion as a biosensor for process inhibitors
14.6 Outlook
References
15 Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater
Chapter Outline
15.1 Introduction
15.1.1 History of beer
15.1.2 Brewing process and wastewater
15.1.3 Brewery waste and beer wastewater treatment
15.1.3.1 Physical treatment
15.1.3.2 Chemical treatment processes
15.1.3.3 Biological treatment methods
15.1.4 Bioelectrochemical systems for beer wastewater treatment
15.1.5 Anaerobic digestion of beer wastewater treatment
15.1.6 Hydrogen production in anaerobic reactors with beer wastewater
15.2 Integrated microbial electrolysis–anaerobic digestion for beer wastewater treatment
15.2.1 Background
15.2.1.1 China-Global beer hub
15.2.1.2 Significance and application prospects of the reactor
15.2.1.3 Unique advantages of microbial electrolysis–anaerobic digestion reactor over conventional technologies
15.2.1.3.1 A complete treatment of a wide range of wastewaters
15.2.1.3.2 Inexpensive upgrading process
15.2.1.3.3 Maintenance of reactor stability
15.2.1.3.4 Hydrogen production increases the speed of methane production
15.2.1.4 Working principle of the reactor
15.2.2 An experience of scaling up of the novel microbial electrolysis–anaerobic digestion reactor
15.2.2.1 Determination of the appropriate cathode electrode material
15.2.2.1.1 Reactor construction and operation
15.2.2.1.2 Sampling and electrochemical analyses
15.2.2.1.3 Outcomes and substantiations
15.2.2.2 Estimation of electrode positions and hydraulic retention time
15.2.2.2.1 Reactor construction and operation
15.2.2.2.2 Electrochemical analyses
15.2.2.2.3 Outcomes and substantiations
15.2.2.3 Estimation of cathode/anode ratio
15.2.2.3.1 Reactor construction and operation
15.2.2.3.2 Electrochemical analysis
15.2.2.3.3 Performance
15.2.3 Overall summary of the experience
15.2.3.1 Cathode selection
15.2.3.2 Electrode positions and hydraulic retention time study
15.2.3.3 Cathode/anode ratio study
Acknowledgments
References
Part 4: Large-scale MFC: potentials and challenges
16 Recent advancements in scaling up microbial fuel cells
Chapter Outline
16.1 Introduction
16.2 Microbial fuel cell designs used in scale-up studies
16.2.1 Larger laboratory reactors
16.2.2 Pilot-scale tests
16.3 Engineering parameters affecting scale-up
16.3.1 Reactor configuration
16.3.2 Internal currents
16.3.3 Membranes
16.3.4 Tubing and compartments
16.4 Design limitations determined by wastewater application
16.4.1 Effect of buffer capacity
16.4.2 Influence of membrane separator
16.4.3 Design limitations determined by scale-up
16.4.3.1 Scale-up and voltage loss
16.4.3.2 Hydrodynamics and mechanics
16.5 Overcoming design constraints
16.6 Life cycle assessment
16.7 Current challenges and potential opportunities
16.8 Conclusion
References
Index
Back Cover

Citation preview

Integrated Microbial Fuel Cells for Wastewater Treatment

Integrated Microbial Fuel Cells for Wastewater Treatment Rouzbeh Abbassi School of Engineering, Faculty of Science and Engineering, Macquarie University, Sydney, NSW, Australia

Asheesh Kumar Yadav Environment and Sustainability Department, CSIR-Institute of Minerals and Materials Technology, Bhubaneswar, India

Faisal Khan Centre for Risk, Integrity and Safety Engineering, Faculty of Engineering and Applied Science, Memorial University of Newfoundland, St. John’s, NL, Canada

Vikram Garaniya Australian Maritime College, College of Sciences and Engineering, University of Tasmania, Launceston, TAS, Australia

Butterworth-Heinemann is an imprint of Elsevier The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, United Kingdom 50 Hampshire Street, 5th Floor, Cambridge, MA 02139, United States Copyright © 2020 Elsevier Inc. All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Details on how to seek permission, further information about the Publisher’s permissions policies and our arrangements with organizations such as the Copyright Clearance Center and the Copyright Licensing Agency, can be found at our website: www.elsevier.com/permissions. This book and the individual contributions contained in it are protected under copyright by the Publisher (other than as may be noted herein). Notices Knowledge and best practice in this field are constantly changing. As new research and experience broaden our understanding, changes in research methods, professional practices, or medical treatment may become necessary. Practitioners and researchers must always rely on their own experience and knowledge in evaluating and using any information, methods, compounds, or experiments described herein. In using such information or methods they should be mindful of their own safety and the safety of others, including parties for whom they have a professional responsibility. To the fullest extent of the law, neither the Publisher nor the authors, contributors, or editors, assume any liability for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions, or ideas contained in the material herein. British Library Cataloguing-in-Publication Data A catalogue record for this book is available from the British Library Library of Congress Cataloging-in-Publication Data A catalog record for this book is available from the Library of Congress ISBN: 978-0-12-817493-7 For Information on all Butterworth-Heinemann publications visit our website at https://www.elsevier.com/books-and-journals

Publisher: Matthew Deans Editorial Project Manager: Isabella C. Silva Production Project Manager: Nirmala Arumugam Cover Designer: Matthew Limbert Typeset by MPS Limited, Chennai, India

Contents

List of contributors

xi

Part 1

1

1

2

Introduction to microbial fuel cells: challenges and opportunities Rouzbeh Abbassi and Asheesh Kumar Yadav 1.1 Introduction 1.2 Brief history of microbial fuel cells to bioelectrochemical systems 1.3 Principles and challenges of microbial fuel cells 1.4 Future of microbial fuel cells 1.5 Conclusion Acknowledgment References Microbial fuel cell integrated wastewater treatment systems Pratiksha Srivastava, Rouzbeh Abbassi, Asheesh Kumar Yadav, Vikram Garaniya and Faisal Khan 2.1 Introduction 2.2 Conclusion References

Part 2 3

Introduction

Application to Industrial Wastewater treatment

Removal of heavy metals using bioelectrochemical systems Sukrampal, Rohit Kumar and Sunil A. Patil 3.1 Introduction 3.2 Bioelectrochemical systems for heavy metal removal 3.3 Electrode materials used for heavy metal removal in bioelectrochemical systems 3.4 Conventional technologies versus bioelectrochemical systems-based technology for the removal of heavy metals 3.5 Conclusion References Further reading

3 3 4 6 14 18 18 18 29

29 41 42

47 49 49 56 58 63 66 67 71

vi

4

5

6

7

Contents

Textile wastewater treatment using microbial fuel cell and coupled technology: a green approach for detoxification and bioelectricity generation Supriya Gupta, Yamini Mittal, Prashansa Tamta, Pratiksha Srivastava and Asheesh Kumar Yadav 4.1 Microbial fuel cell and its application in the treatment 4.2 Enhancement of microbial fuel cell performance 4.3 Microbial diversity involved in the breakdown of dye in microbial fuel cell 4.4 Toxicity of treated dye wastewater 4.5 Microbial fuel cell coupled techniques for textile wastewater treatment 4.6 Research gap Acknowledgments References Further reading Agro-industrial wastewater treatment in microbial fuel cells Silvia Bolognesi, Daniele Cecconet and Andrea G. Capodaglio 5.1 Introduction 5.2 Use of agro-industrial wastewater as substrate for microbial fuel cells 5.3 Dairy industry wastewater 5.4 Brewery and winery industry 5.5 Agro-industrial wastewaters and by-products 5.6 Livestock industry wastewater 5.7 Challenges in using microbial fuel cells 5.8 Conclusion References Pharmaceutical wastewater treatment in microbial fuel cell Somdipta Bagchi and Manaswini Behera 6.1 Introduction 6.2 Application to pharmaceutical wastewater treatment 6.3 Integration of microbial fuel cell with other wastewater-treatment processes 6.4 Large-scale microbial fuel cell: potentials and challenges References Oil and petrochemical industries wastewater treatment in bioelectrochemical systems Surajbhan Sevda, Vijay Kumar Garlapati, Swati Sharma, Udaratta Bhattacharjee, Lalit Pandey and T.R. Sreekrishnan 7.1 Introduction 7.2 Oil field and petrochemical wastewater treatment in the conventional treatment process

73

74 78 81 82 83 86 87 87 91 93 93 95 96 101 107 117 120 123 123 135 135 138 146 147 148

157

157 158

Contents

Oil field and petrochemical wastewater treatment in the bioelectrochemical system 7.4 Conclusion References Further reading

vii

7.3

8

9

Bioelectrochemical systems for stormwater treatment and energy valorization processes Kiran Tota-Maharaj and Marika E. Kokko 8.1 Introduction 8.2 Energy recovery and stormwater treatment efficiency with bioelectrochemical systems 8.3 Economic and environmental considerations 8.4 Technical scales of bioelectrochemical systems 8.5 Outlook, challenges, and future perspectives 8.6 Conclusion References Further reading Treatment of food processing and beverage industry wastewaters in microbial fuel cells Abhilasha Singh Mathuriya and Soumya Pandit 9.1 Introduction 9.2 Food-based wastes and wastewater as a substrate for microbial fuel cell 9.3 Beer brewery wastewater wastes and wastewater as a substrate for microbial fuel cell 9.4 Conclusion References Further reading

Part 3 Integration of MFC with other wastewater treatment processes 10

Microbial fuel cell coupled with microalgae cultivation for wastewater treatment and energy recovery P. Chiranjeevi and Sunil A. Patil 10.1 Introduction 10.2 Microbial fuel cell and microalgae cultivation based integrated systems 10.3 Factors influencing the performance of integrated microbial fuel cell and microalgae cultivation systems 10.4 Conclusion Acknowledgment References

164 168 169 173

175 175 179 185 187 189 191 192 197

199 199 200 202 203 206 210

211 213 213 215 221 223 223 224

viii

11

12

13

14

Contents

Integration of bioelectrochemical systems with other existing wastewater treatment processes M. M. Ghangrekar, Sovik Das and Bikash R. Tiwari 11.1 Introduction 11.2 Integration of bioelectrochemical system with electro-Fenton process 11.3 Integration of bioelectrochemical system with aerobic processes 11.4 Integration of microbial fuel cell with anaerobic digestion 11.5 Microbial fuel cell integration with septic tank 11.6 Microbial fuel cell integration with dark fermentation 11.7 Microbial fuel cell integration with microalgae 11.8 Novel integration of other processes with microbial fuel cells 11.9 Way forward References An overview of membrane bioreactor coupled bioelectrochemical systems M. M. Ghangrekar, G.D. Bhowmick and S.M. Sathe 12.1 Introduction 12.2 Bioelectrochemical systems 12.3 Membrane bioreactor 12.4 Hybrid bioelectrochemical system membrane bioreactor systems: principle, treatment efficiency, and performance index 12.5 Outlook and future perspectives 12.6 Conclusion Acknowledgment References Integration of microbial fuel cell into constructed wetlands: effects, applications, and future outlook Supriya Gupta, Pratiksha Srivastava and Asheesh Kumar Yadav 13.1 Introduction 13.2 Development of merger technology 13.3 Challenges and future perspectives Acknowledgment References Microbial fuel cell coupled with anaerobic treatment processes for wastewater treatment Suman Bajracharya 14.1 Introduction 14.2 Integration of microbial fuel cell in anaerobic digestion 14.3 Microbial fuel cell coupling to treat undigested organics in the effluents of anaerobic digestion

229 229 232 235 240 241 242 242 243 244 245

249 249 251 252 254 262 267 267 268

273 273 278 287 288 288

295 295 297 299

Contents

ix

14.4

Microbial fuel cells coupled anaerobic digestion for nutrient recovery and toxicity removal 14.5 Microbial fuel cell coupling in anaerobic digestion as a biosensor for process inhibitors 14.6 Outlook References 15

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater Thangavel Sangeetha, Chellappan Praveen Rajneesh and Wei-Mon Yan 15.1 Introduction 15.2 Integrated microbial electrolysis anaerobic digestion for beer wastewater treatment Acknowledgments References

Part 4 16

Large-scale MFC: potentials and challenges

Recent advancements in scaling up microbial fuel cells Soumya Pandit, Nishit Savla and Sokhee P. Jung 16.1 Introduction 16.2 Microbial fuel cell designs used in scale-up studies 16.3 Engineering parameters affecting scale-up 16.4 Design limitations determined by wastewater application 16.5 Overcoming design constraints 16.6 Life cycle assessment 16.7 Current challenges and potential opportunities 16.8 Conclusion References

Index

303 304 305 306

313

313 319 342 342

347 349 349 350 355 357 360 362 363 364 364 369

List of contributors

Rouzbeh Abbassi School of Engineering, Faculty of Science and Engineering, Macquarie University, Sydney, NSW, Australia Somdipta Bagchi School of Infrastructure, IIT Bhubaneswar, Bhubaneswar, India Suman Bajracharya Water Desalination and Reuse Center, King Abdullah University of Science and Technology, Thuwal, Saudi Arabia Manaswini Behera School of Infrastructure, IIT Bhubaneswar, Bhubaneswar, India Udaratta Bhattacharjee Center for the Environment, Indian Institute of Technology Guwahati, Guwahati, India G.D. Bhowmick Department of Agricultural and Food Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India Silvia Bolognesi Department of Civil Engineering and Architecture, University of Pavia, Pavia, Italy Andrea G. Capodaglio Department of Civil Engineering and Architecture, University of Pavia, Pavia, Italy Daniele Cecconet Department of Civil Engineering and Architecture, University of Pavia, Pavia, Italy P. Chiranjeevi Department of Earth and Environmental Sciences, Indian Institute of Science Education and Research Mohali (IISER Mohali), Sector 81, S.A.S. Nagar, Manauli, Punjab, India Sovik Das Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India Vikram Garaniya Australian Maritime College, College of Sciences and Engineering, University of Tasmania, Launceston, TAS, Australia Vijay Kumar Garlapati Department of Biotechnology and Bioinformatics, Jaypee University of Information Technology (JUIT), Waknaghat, India

xii

List of contributors

M.M. Ghangrekar Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India Supriya Gupta Academy of Scientific and Innovative Research (AcSIR), CSIRHuman Resource Development Centre, (CSIR-HRDC) Campus, Ghaziabad, India; Environment and Sustainability Department, CSIR-Institute of Minerals and Materials Technology, Bhubaneswar, India Sokhee P. Jung Department of Environment and Energy Engineering, Chonnam National University, Gwangju, South Korea Faisal Khan Centre for Risk, Integrity and Safety Engineering, Faculty of Engineering and Applied Science, Memorial University of Newfoundland, St. John’s, NL, Canada Marika E. Kokko Bio and Circular Economy Research Group, Department of Chemistry and Bioengineering, Tampere University, Tampere, Finland Rohit Kumar Department of Earth and Environmental Sciences, Indian Institute of Science Education and Research (IISER Mohali), Mohali, Punjab, India Abhilasha Singh Mathuriya Department of Life Sciences, School of Basic Sciences and Research, Sharda University, Greater Noida, India Yamini Mittal Environment and Sustainability Department, CSIR-Institute of Minerals and Materials Technology, Bhubaneswar, India Lalit Pandey Department of Bioscience and Biotechnology, Indian Institute of Technology Guwahati, Guwahati, India Soumya Pandit Department of Life Sciences, Sharda University, Greater Noida, India Sunil A. Patil Department of Earth and Environmental Sciences, Indian Institute of Science Education and Research (IISER Mohali), Mohali, Punjab, India Chellappan Praveen Rajneesh School of Biomedical Engineering, College of Biomedical Engineering, Taipei Medical University, Taipei, Taiwan, ROC Thangavel Sangeetha Department of Energy and Refrigerating Air-Conditioning Engineering, National Taipei University of Technology, Taipei, Taiwan, ROC; Research Center of Energy Conservation for New Generation of Residential, Commercial, and Industrial Sectors, National Taipei University of Technology, Taipei, Taiwan, ROC

List of contributors

xiii

S.M. Sathe Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India Nishit Savla Amity Institute of Biotechnology, Amity University, Mumbai, India Surajbhan Sevda Department of Bioscience and Biotechnology, Indian Institute of Technology Guwahati, Guwahati, India; Department of Biotechnology, National Institute of Technology Warangal, Warangal, India Swati Sharma Department of Biotechnology and Bioinformatics, University of Information Technology (JUIT), Waknaghat, India

Jaypee

T.R. Sreekrishnan Department of Biochemical Engineering and Biotechnology, Indian Institute of Technology Delhi, New Delhi, India Pratiksha Srivastava Australian Maritime College, College of Science and Engineering, University of Tasmania, Launceston, TAS, Australia Sukrampal Department of Earth and Environmental Sciences, Indian Institute of Science Education and Research (IISER Mohali), Mohali, Punjab, India Prashansa Tamta University School of Environment Management, Guru Gobind Singh Indraprastha University, New Delhi, India Bikash R. Tiwari Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India Kiran Tota-Maharaj Civil & Environmental Engineering Cluster, Faculty of Environment and Technology, Centre for Water, Communities and Resilience & The International Water Security Network, University of the West of England, Bristol (UWE Bristol), Frenchay Campus, Bristol, United Kingdom; Bio and Circular Economy Research Group, Department of Chemistry and Bioengineering, Tampere University, Tampere, Finland Asheesh Kumar Yadav Academy of Scientific and Innovative Research (AcSIR), CSIR-Human Resource Development Centre, (CSIR-HRDC) Campus, Ghaziabad, India; Environment and Sustainability Department, CSIR-Institute of Minerals and Materials Technology, Bhubaneswar, India Wei-Mon Yan Department of Energy and Refrigerating Air-Conditioning Engineering, National Taipei University of Technology, Taipei, Taiwan, ROC; Research Center of Energy Conservation for New Generation of Residential, Commercial, and Industrial Sectors, National Taipei University of Technology, Taipei, Taiwan, ROC

Introduction to microbial fuel cells: challenges and opportunities

1

Rouzbeh Abbassi1 and Asheesh Kumar Yadav2 1 School of Engineering, Faculty of Science and Engineering, Macquarie University, Sydney, NSW, Australia, 2Academy of Scientific and Innovative Research (AcSIR), CSIRHuman Resource Development Centre, (CSIR-HRDC) Campus, Ghaziabad, India; Environment and Sustainability Department, CSIR-Institute of Minerals and Materials Technology, Bhubaneswar, India

Chapter Outline 1.1 Introduction 3 1.2 Brief history of microbial fuel cells to bioelectrochemical systems 1.3 Principles and challenges of microbial fuel cells 6 1.4 Future of microbial fuel cells 14 1.5 Conclusion 18 Acknowledgment 18 References 18

1.1

4

Introduction

Contaminated drinking water, inappropriate sanitation services, and inadequate handwashing facilities were the cause of 800,000 deaths worldwide in 2012 [WWAP (United Nations World Water Assessment Programme), 2017]. Moreover, the discharge of the untreated wastewaters into different water bodies affected approximately 245,000 km2 of marine ecosystems that impacted fisheries and other food chains [United Nations Educational, Scientific and Cultural Organization (UNESCO), 2017]. Despite the large volume of wastewater produced daily being a threat to the environment, it contains nutrient that is valuable for agriculture. The scarcity of natural resources of nutrients and water has dramatically increased the need to recognize the importance of wastewater. The large volume of treated water is used for different purposes, including for irrigation (Angelakis and Snyder, 2015). Certain nutrients such as phosphorus and nitrogen, abundantly found in wastewaters (Abbassi et al., 2014; Egle et al., 2016), can be recovered and converted to fertilizer (Sengupta et al., 2015; Gunther et al., 2018). Recovering energy from the wastewater treatment process leads to many environmental benefits. The reductions in greenhouse gas emissions and air pollution are just some of the side benefits of energy recovery from wastewater (Smith et al., 2018). A major drawback most of the wastewater treatment systems is the high energy consumption. The cost associated with the energy is between 2% and 60% of the Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00001-1 © 2020 Elsevier Inc. All rights reserved.

4

Integrated Microbial Fuel Cells for Wastewater Treatment

total operating cost of a wastewater treatment plant (Mizuta and Shimada, 2010; Guerrini et al., 2017; Smith and Liu, 2017). The existing potential chemical energy stored in wastewater is much higher than the energy required for treating the wastewater (Khiewwijit et al., 2015). There is an urgent need to develop the future generation of wastewater treatment plants, which should have energy-neutral operation, capable of the recovery of energy and nutrients such as nitrogen and phosphorous (Puyol et al., 2017). At the same time, microbial fuel cells (MFCs) are upcoming technology that is capable of generating electricity by oxidizing a variety of pollutants from wastewater by using microorganisms. MFCs can convert energy stored in chemical bonds in biodegradable compounds to electrical energy. Although the MFCs are not yet ready to replace the conventional fossil-fuelbased energy generation, these may be an alternative source of energy in the near future (Slate et al., 2019). This chapter provides a brief overview of the current understanding of MFCs and allied technologies and discusses the potential applications in the field of wastewater treatment in addition to electricity generation.

1.2

Brief history of microbial fuel cells to bioelectrochemical systems

The concept of MFC is not new. Michael C. Potter demonstrated the concept in 1910
11 using microbial induced electrode reduction with Escherichia coli cultures (Potter, 1911). In his experiment, electricity achieved was not significantly usable, so little attention was given to the developed concept. Cohen (1931), and later Davis and Yarborough (1962), created the first proper MFC. Cohen drew further attention to MFCs by presenting microbial half fuel cells that generated 35 V and a current of about 0.2 mA (Cohen, 1930). In 1963 the microbial electricity generation peaked again by presenting the idea of utilization of human wastes as electrochemical fuels by the National Aeronautics and Space Administration (NASA) space program (Canfield et al., 1963). In the early 1990s, researchers (Habermann and Pommer, 1991) reported that instead of synthetic mediators, electrochemically active bacteria could produce natural shuttle molecules for electron transfer. Their findings have been a pioneer for treating wastewater using MFCs. Since then, using MFCs to harness energy from the metabolism of microorganisms has become attractive for energy generation and wastewater treatment (Logan, 2004; Logan and Regan, 2006b; Watanabe, 2008; Luo et al., 2010; Koroglu et al., 2014; Gajda et al., 2018). Although most attempts until now have been around the application of MFCs for electricity generation and wastewater treatment in the laboratory spaces, however, attempts have been rarely implemented in practice and real-life environments. A century of knowledge of the MFCs’ fundamentals does still not guarantee their long-term practical applications for energy generation. Water and wastewater treatment aspects in MFCs have matured and are nearly ready for larger level implementation. In recent years, due to extensive interest and constant new applications with MFCs, they have extended into bioelectrochemical systems (BESs). Still, MFCs are

Introduction to microbial fuel cells: challenges and opportunities

5

Figure 1.1 Schematic of a microbial fuel cell (A), microbial electrolysis cell (B), microbial desalination cell (C), and general microbial electrosynthesis cell (D). Source: Santoro, C., Arbizzani, C., Erable, B., Ieropoulos I., 2017. Microbial fuel cells: from fundamentals to applications
a review. J. Power Sources 356, 225
244, Taken under Creative Commons Attribution License.

the most studied and reported BESs (Santoro et al., 2017). Fig. 1.1 provides a general schematic diagram of MFCs along with other BESs. Besides MFCs, other BESs that are known as microbial electrolysis cells (MECs), are investigated with huge interest. MEC was first introduced in 2005 (Liu et al., 2005) and is becoming popular due to their capabilities to produce hydrogen which is an upcoming hydrogen economy (Santoro et al, 2017). The operation of MECs needs a small amount of external electric energy in addition to self-generated electric energy to produce hydrogen at the cathode (Sleutels et al., 2009). As a result, hydrogen can be formed in MECs with a low consumption of energy utilizing bioelectrocatalysis supported by additional low energy power sources. In recent years, large-scale applications of MFCs have been presented such as production of hydrogen from winery wastewater (Cusick et al, 2011), and pilotscale application of MEC using urban wastewater (Baeza et al., 2017). These types of applications with MECs plainly indicate that the MECs are evolving toward practical applications. Recent various other BESs are also coming into the picture with new applications such as the microbial desalination cell (MDC) that was developed for desalinating water and wastewater, and to simultaneously generate electricity. A 105 L capacity MDC system was recently presented by Zhang and He (2015). More recently, more exciting versions of BESs have been introduced as microbial electrosynthesis cells for synthesizing valuable products at cathode from CO2

6

Integrated Microbial Fuel Cells for Wastewater Treatment

or other compounds from gas transformation or reduction. This direction of BESs is also gaining high interest due to the possible utilization of renewable energy. The main principles of the MECs are shown in Fig. 1.1. Microbial electrosynthesis cells can convert CO2 to methane (Rabaey and Rozendal, 2010), acetate (Rabaey and Rozendal, 2010; Rabaey et al., 2011), formate, and other compounds (Roy et al., 2016; Huang et al., 2016).

1.3

Principles and challenges of microbial fuel cells

MFCs are capable of converting the chemical energy in organic matter to electricity by adopting metabolic processes of microorganisms (Chouler and Lorenzo, 2015). Therefore a dual advantage of wastewater treatment and electricity generation is expected in MFCs. The original MFCs are designed in two chambers, one is anode and other is cathode. These two chambers are separated by a membrane. Organic matters act as fuel in the anode compartment (Ucar et al., 2017). In the anode the bacteria consume organic matters in the wastewater through metabolism (anaerobic degradation) and generate electrons and protons. An oxidation reaction at the anode for acetate as an example is illustrated in Reaction (1.1) (Pandit et al., 2017). CH3 COO2 1 2H2 O ! 2CO2 1 7H1 1 8e2

(1.1)

In the cathode chamber an electron acceptor uses electrons. Oxygen is one of the best known electron acceptors used in MFCs as it has high oxidation potential, and it becomes a clean product (water) after reduction. The transferred electrons thus combine with protons and oxygen in the cathode compartment to produce water, as explained in Reaction (1.2). 2O2 1 8H1 1 8e2 ! 4H2 O

(1.2)

In recent years, other electron acceptors such as ferricyanide, nitrate, permanganate, persulfate, ferric iron, mercury, and chromium have also been tested and reported to be used in the cathode compartment (Heijne et al., 2006; Li et al., 2009; Lu and Li, 2012; Oh et al., 2004; You et al., 2006; Wang et al., 2011; Zhao et al., 2016). Electron transfer is from anode to cathode by using an external circuit, while protons migrate between the two chambers by passing through a proton-exchange membrane (PEM). PEM is an important component of the MFC as it facilitates the transfer of protons to the cathode chamber to have a sustainable electric current and charge neutrality in the system (Tharali et al., 2016). An optimal PEM should avoid the transfer of oxygen, substances, and minerals between the chambers (Rahimnejad et al., 2014). Nafion is one of the most common PEMs used in MFCs (Chae et al., 2008). The cost of common PEMs such as Nafion is one of the downsides of using those in MFCs (Flores et al., 2015). The cost of PEMs is reported to

Introduction to microbial fuel cells: challenges and opportunities

7

be approximately 40% of the total cost of a finished MFC (Rozendal et al., 2008). The transport of cations other than protons and biofouling of the Nafion membranes are the other challenges (Daud et al., 2015).Therefore replacing the Nafion by other membranes such as nylon and glass fiber filters, earthen pot, porous ceramic, composite and nanocomposite membranes, or using membraneless MFCs are the alternative options to optimize the power generation at low cost (Ghasemi et al., 2012; Daud et al., 2015; Moon et al., 2006; Zhuwei et al., 2008). Considering the challenges of membranes (separators) such as increased porosity, proton conductivity, brittleness, and cost, investigating the new materials utilized as PEMs is one the future needs for commercializing the MFCs operations. The performance of the MFCs is heavily dependent on the electrode materials. The electrode materials affect bacterial adhesion, electron transfer, and electrochemical efficiency (Mustakeem, 2015). Biocompatibility, chemical stability, anticorrosiveness, sufficient mechanical strength, and toughness are other factors that are also important to be considered for anode electrode materials (Zhou et al., 2011). Carbon materials such as graphite plates, graphite rod, carbon cloth, carbon paper, graphite fiber brush, and carbon felt are common electrode materials used in MFCs (Logan et al., 2007; Zhao et al., 2018). Carbon nanotubes (CNTs) are reported as promising materials used for MFC electrodes (Yazdi et al., 2016). Recently, composite materials such as polypyrrole
CNT showed the capacity for a high electron transfer. Adding the carbon materials such as graphite granules, activated carbon granules, and activated carbon powder to the cathode chamber of MFCs enhances the power generation (Tursun et al., 2016). Other conductive materials such as stainless steel, gold sheet, silver sheet, and titanium plate have also been utilized as electrode materials in MFCs (Pocaznoi et al., 2012; Santoro et al., 2017). Some of the known electrode materials used in MFCs are depicted in Table 1.1. To enhance the performance of different types of MFCs, not only are the electrode materials important but also the positions and configurations of the electrodes placed in reactors are also very important (Pushkar et al., 2018). Overall, exploration of new cost-effective electrode materials to enhance the MFCs efficiency in both power generation and wastewater treatment is among significant areas where more attention is needed for widening the application and commercialization of MFCs. The transfer of electrons to anodes often occurs directly through microbial membrane, electron mediators, or shuttles or electrically conductive appendages called nano-wires (Logan, 2009). Direct electron transfer may occur by physical contact between the bacterial cell membrane and the electrode (anode) (Schroder, 2007). In some cases the electrons produced by the microorganisms during oxidation of substrate need electron shuttles of the mediator to be transferred to the electrode (Fultz and Durst, 1982). A mediator is a compound having low redox potential that is added to extract of the electrons from the metabolic reactions of the microbes in and supplies those electrons to the anode electrode (Sevda and Sreekrishnan, 2012). Electron shuttles (mediators) are low molecular weight organic molecules that are capable of catalyzing reduction/oxidation reactions (Velasquez-Orta et al., 2010). The electrons produced during anaerobic oxidation of the substrate by the bacteria

Table 1.1 Characteristics of different electrode materials used in microbial fuel cells (MFCs). Electrode material

Type of MFCs

Key features

Anode

Cathode

Graphite plate

Stainless steel wire

Sediment

G

Carbon paper

Carbon paper containing Pt catalyst Three graphite rods

Two chamber

G

Stainless steel

G

Dual chamber

Powering wireless sensor up to 320 mV Power density of 239.4 mW/m2 Current density of 893.3 mA/m2

Graphite foil

Donovan et al. (2008) Lu et al. (2009) Jadhav and Ghangrekar (2009)

Multiple stack

With higher COD removal (90%), low current production (0.7 mA), and low CE (1.5%) With low COD removal (59%), higher current production (1.4 mA) and CE (4%) Power density of 0.80 W/m2 and current density of 3.2 A/m2 Pure copper crystal recovery at cathode Power density of 1.24 W/m2 Volumetric power density of 182 W/m3 2.5 times higher power density and 12 times higher volumetric density from the previous reported studies COD removal of 83.8% and NH41
N
90.8%

Dual chamber

Current of 1.9 6 0.4 mA and power density of 6.0 W/m3

Wang et al. (2012)

G

G

Rough (sandpapered) graphite plate CNT sponge

Reference

Dual chamber

G

Heijne et al. (2010)

G

CNT sponge-Pt

Dual chamber

G

G

Xie et al. (2012)

G

Graphite felt

Activated carbon fiber

Hot pressed carbon fiber cloth containing MnO2 catalyst (0.8 mg/cm2) Carbon felt

Zhuang et al. (2012)

Carbon felt coated with ruthenium oxide Carbon cloth coated with graphene oxide and polyaniline 3D reduced graphene oxide
nickel Carbon brushes

Graphite fiber brush

GAC, GGs, BCp, BCc

Pt-mesh

Dual chamber

G

G

Carbon felt-Pt coated

Dual chamber

Carbon cloth

Dual chamber

G

G

Based on anode electrode material, volumetric power density of 661 W/m3 Based on anode volume, power density of 27 W/m3 Current of 0.435 6 0.01 A, and power density of 116 mW, COD removal of 79% 6 7%, and TN removal of 17% 6 8% Total COD removal of 92.5% and complete removal of TSS

G

Lv et al. (2012) Hou et al. (2013)

Wang et al. (2013)

G

Carbon mesh loaded with Pt

Stacked

Carbon cloth

Two- and four-stage MFC, hydraulically connected in series combined with anaerobic fluidized bed membrane reactor Dual chamber

Carbon cloth

G

G

G

Graphite fiber brush

Power density of 3.08 W/m2 17 times high power density than the bare anode MFC Graphene act as highly conductive support material Charge transfer rate and bacterial biofilm loading enhanced

Poly binder with a mixture of activated carbon and carbon black

Single chamber

G

G

Power production of BCp (532 6 18 mW/m2), and BCc (457 6 20 mW/m2) comparable with GAC (674 6 10 mW/m2) and GG (566 6 5 mW/m2) Low material expense made power output cost of 17
35 US$/W, 90% cheaper than GAC (402 US $/W) or GG (392 US$/W) Two cathodes produced more power than single cathode Single-cathode MFCs removed higher COD (54.2% 6 2.3%, N1C) than the two-cathode MFCs (48.3% 6 1.0%, S2C)

Feng et al. (2014) Ren et al. (2014)

Huggins et al. (2014)

Kim et al. (2015)

(Continued)

Table 1.1 (Continued) Electrode material

Type of MFCs

Key features

Anode

Cathode

3D GA
Pt nanoparticle 3D-carbon black/ stainless steel mesh Copper, silver

Pt sheet

Dual chamber

G

Graphite plate

Dual chamber

G

Carbon brush

Carbon paper

G

G

Graphite rod

Single chamber

G

705 wt high surface area activated carbon, 10 wt.% carbon black and 20 wt.% PTFE Pani
MnO2

Single chamber

G

G

G

Dual chamber

G

G

AA:Ni
ACF/CNFs

AA:Ni
ACF/CNFs

Mediator less double chamber

G

G

Flat carbon cloth

Carbon cloth

Floating MFC

G

G

Reference

Power density of 1460 mW/m2, 5.3 times higher than carbon cloth (273 mW/m2) Current density of 10.07 6 0.88 mA/cm2 Volumetric current density of 18.66 6 1.63 mA/cm3 Maximum power density of 3215 6 80 mW/m2 Current density of 1.1 mA/cm2 (silver) and 1.5 mA/ cm2 (copper) Power production of 19 mW Geometric cathode area of 84 W/m2 Highest power ever achieved in MFCs

Zhao et al. (2015) Zheng et al. (2015)

Power density of 0.0588 W/m2 The fibrous Pani
MnO2 nanocomposite achieved high capacitance (525 F g/1 at a current density of 2 A g/1) and excellent cycling stability of 76.9% after 1000 cycles at 10 A g/1 Complete removal of Cr(VI) at 100 ppm concentration Maximum power density of B1540 mW/m2 Power production of 3
3.5 mW (power density of 22
28 mW/m2 cathode) The average of daily electrical energy harvested ranged between 10 and 35 mWh/d

Ansari et al. (2016)

Baudler et al. (2015) Santoro et al. (2016)

Gupta et al. (2017) Schievano et al. (2017)

Activated carbon fiber felt

Activated carbon fiber felt

Soil based, single chamber

G

G

G

G

Screen printed, conductive ink
based biodegradable, carbon-based paper electrode Graphite brush

Screen printed, conductive ink
based biodegradable, carbon-based paper electrode Carbon cloth-Pt

Carbon brush

FE-AAPyr, graphene nanosheets Manganese-based catalyzed carbon

Carbon felt

Plant

G

Air cathode, two chamber

G

Dual chamber

G

G

Air cathode, single chamber

G

G

Graphite fiber brush

Stainless steel sheet

Large scale

G

G

Power density of 12.1 mW/m2 Anthracene removal of 54.2% 6 2.7% Phenanthrene removal of 42.6% 6 1.9% Pyrene removal of 27.0% 6 2.1% A successful biosensor developed for pollutant monitoring

Yu et al. (2017)

Complete removal of Cr(VI) from 80 mg/L Cr(VI) in carbon cloth cathode Power generation of 235 6 1 µW/cm2 Higher performance than graphene nanosheets

Farahani et al. (2018) Kodali et al. (2018)

COD removal of 90%, ammonia removal of 98%, and total nitrogen removal of 95% Maximum power output 1250 6 20 mW/m2 HRT of 22 min improved 17% performance of power density, 0.118 6 0.006 W/m2 Reduction of diameter of anode brush did not improve the performance

Yang et al. (2019)

Chouler et al. (2018)

Rossi et al. (2019)

AA, Alumina; ACF, activated carbon microfiber; BCc, biochar of forestry residue; BCp, biochar of compressed milling residue; CE, coulombic efficiency; CNFs, carbon nanofibers; CNT, carbon nanotube; FE-AAPyr, iron aminoantipyrine; GA, graphene aerogel; GAC, granular activated carbon; GGs, graphite granules; Ni, nickel; COD, chemical oxygen demand; TSS, total suspended solids; TN, total nitrogen; PTFE, polytetrafluoroethylene; BOD, biochemical oxygen demand; DO, dissolved oxygen; TCOD, total chemical oxygen demand; SCOD, soluble chemical oxygen demand.

12

Integrated Microbial Fuel Cells for Wastewater Treatment

are transferred to the anode either directly from the enzyme involved in bacterial respiration to the electrode either with or without (Bond and Lovley, 2003; Chaudhuri and Lovely, 2003; Gil et al., 2003) the help of mediators or through an external electron carrier compound (potassium ferricyanide, thionine, neutral red, methylene blue) (Delaney et al., 1984; Park and Zeikus, 2000; Rabaey et al., 2004a,b; Babanova et al., 2011). External or exogenous chemical compounds used as mediators in MFC have certain drawbacks such as toxicity and high cost (Bond and Lovley, 2003; Gil et al., 2003). An ideal mediator present in an MFC should have specific characteristics such as (Park and Zeikus, 2000) (1) solubility in aqueous systems, (2) capability of forming a reversible redox couple at the electrode, (3) must be stable in both oxidized and reduced form, and (4) capability of linking to NADH and have a high negative E02 value. The mediators need to be stable in both reduced and oxidized conditions, should not be biologically degradable, and should not be toxic for microorganisms (Mahadevan et al., 2014). Different mediators such as methylene blue, neutral red, thionine, and potassium ferricyanide have been previously tested in MFCs to enhance their performances (Park and Zeikus, 2000; Rahimnejad et al., 2011; Pham et al., 2004). The electrical circuit elements, particularly wires, connected to electrodes are another factor affecting the MFCs performance. They directly affect the electron transfer and power generation in MFCs. Factors such as higher conductivity, lower resistivity, corrosion resistance, and cost should be considered in selecting an appropriate wire. Different types of electric wires such as titanium, copper, stainless steel, tungsten, and platinum wires have been tested in MFCs (Idris et al., 2016; Jung et al., 2018; Park and Zeikus, 2000; Sharma and Ghangrekar, 2018). Comparison of tungsten, stainless steel, and titanium wires confirmed a higher power density and current density with smaller internal resistance at MFCs which used tungsten wire as current collector (Sharma and Ghangrekar, 2018). Platinum wires are another option that has been tested in MFCs, and they have high conductivity and low resistivity (Li et al., 2018). However, due to their very high cost, it is not possible to use platinum wire in real-world application. MFCs are divided into different categories on the basis of different components and configurations utilized in their construction. Considering the existence of PEMs, MFCs are divided into two-chamber and single-chamber MFCs (Figs. 1.2 and 1.3, subsequently). The processes associated with two-chamber MFCs are mostly discussed in the abovementioned sections. However, in single-chamber MFCs the anode chamber is linked to a porous air cathode with gas diffusion PEMs separating the chambers (Das and Mangwani, 2010). Applications of singlechamber MFCs for treating different municipal and industrial wastewaters have been reported by different researchers (Marashi and Kariminia, 2015). Although single-chamber MFCs are simpler, more cost effective and produce power in a more efficient way, the priority of two-chamber MFCs is that the performance of the cathode can be enhanced by controlling different factors such as pH, oxygen level, increasing flow rate, and adding electron mediators in the cathode. This leads to enhancement of the performance of MFCs.

Introduction to microbial fuel cells: challenges and opportunities

13

Figure 1.2 A schematic of two-chamber MFCs. MFCs, Microbial fuel cells.

Figure 1.3 A schematic of a single-chamber MFC. MFC, Microbial fuel cell.

Application of MFCs to different types of wastewater confirms their capability to remove the contaminants in both municipal and industrial wastewaters. The MFCs can be used for removing carbon and nitrogen contaminates from municipal wastewater, on the side of electricity generation (Puig et al., 2011; Linares et al., 2019). Using MFCs for treating different types of industrial wastewater from refineries, dairy, food, brewery, landfill leachate, mining, etc. has been reported by different scholars (Addi et al., 2018; Angosto et al., 2015; Ganesh and Jambeck, 2013; Luo et al., 2017; Mansoorian et al., 2016; Ni et al., 2016). Peng et al. (2017) used MFC for removing sulfate and heavy metals from acid mine drainage. Removal of 1298 6 617 mg/L COD from wastewater in diary industry by using continuous MFC operation has been reported (Faria et al., 2017). Jayashree et al. (2015) used MFC for removing the recalcitrant pollutants such as phenol in addition to COD removal. The performance of some of these MFCs applied for removing the

14

Integrated Microbial Fuel Cells for Wastewater Treatment

contaminants from industrial wastewater is depicted in Table 1.2. It should be noted that the capability of the MFCs to degrade a wide range of environmental pollutants (including recalcitrant contaminants) is more valuable than electricity generation itself at this stage, particularly in certain settings, especially when the MFC technology can be used for in situ environmental remediation (Rathoure and Dhatwalia, 2015).

1.4

Future of microbial fuel cells

Developing and utilizing the MFCs at the commercial level still need significant enhancement in different operational sections of this technology to develop a robust system (Mukherjee et al., 2018). The field-scale application of MFCs requires overcoming many technical challenges. The cost of constructing MFCs such as electrode materials and separators is one of the main concerns in large-scale applications (Logan and Regan, 2006a). One of the current challenges of MFCs is their intrinsic intermittent nature (Khera and Chandra, 2012). Developing appropriate devices to store the produced energy is necessary to make the MFC technology more practical. Thus the scale-up of MFCs requires more fundamental research on MFC structure to increase volume and energy output (Jadhav et al., 2018). Reviewing the trend of different studies does not confirm the short-term viability of using large-scale MFCs. However, considering the current capabilities of MFCs, they can be a valuable option for many different applications in water and wastewater engineering. Applications of MFCs for biosensing purposes have received significant attention (Sun et al., 2015). Yang et al. (2015) reviewed different achievements in using MFCs for BOD, DO, volatile fatty acid measurements and also for determining the presence of toxic substances and corrosivity of corrosive biofilms. Integrating the MFCs concept into conventional wastewater treatment processes such as activated sludge, membrane bioreactors, and constructed wetlands is another viable option with different advantages such as increasing the treatment efficiency, sludge reduction and providing a self-sustainable treatment process (Gajaraj and Hu, 2014; Yadav et al., 2012; Xiao et al., 2017). However, due to nonideal operating conditions, high internal resistance, and low columbic efficiency, the power density generated in these hybrid treatment technologies is low (Xu et al., 2016). It therefore seems that the main focus of these integrated systems in the short term should be on presenting more efficient wastewater treatment technologies, not predominantly electricity generation. Although to use the full potential of the integrated wastewater treatment technologies with MFCs in large-scale applications, more focus on MFCs’ construction materials such as developing low-cost electrodes is needed. Another potential application of MFCs is for disinfection purposes of the effluent from the treatment systems. Application of MFCs as a disinfectant/cleaner has been reported previously by researchers (Santoro et al., 2012; Gajda et al., 2016; Singh et al., 2016). There are reports of the use of certain catholytes in MFC, which enhance the power output of MFC but these could have

Introduction to microbial fuel cells: challenges and opportunities

15

Table 1.2 Performance of microbial fuel cells (MFCs) for treating industrial wastewater. Type of wastewater

Type of MFCs

Brewery

Single chamber, air cathode

Beer brewery

Chocolate industry

Single chamber, membrane free Two chamber, activated sludge

Key performances G

G

G

G

G

G

Starch processing industry

Two chamber, air cathode

Food industries: FAJ, wine lees, and YW

Two chamber

G

G

G

G

G

Rice mill

Two chamber, earthen pot

G

G

G

G

Cassava mill Industrial wastewater: bakery, brewery, paper, and dairy Dye

Sediment type Single chamber

Single chamber

G

G

G

References

Temperature has direct effect on power density Addition of 50-mM phosphate buffer increased power output by 136% to 438 mW/m2 Maximum power density of 12 W/m3 at 30 C

Feng et al. (2008)

Current with membrane was 3.02 A/m2 Current with salt bridge was 2.3 A/m2 COD removal of 75% Maximum voltage of 490.8 mV Power density of 239.4 mW/m2 YW, power density of 54 mW/m2 at 232 mA/m2 FAJ, power density of 78 mW/m2 at 209 mA/m2 Wine lees were not suitable Maximum COD removal of 96.5% Lignin removal of 84% Phenol removal of 81% Maximum power density of 2.3 W/m3 Maximum power density of 1771 mW/m2 Higher current density of 125 6 2 mA/m2 achieved from paper industry wastewater

Patil et al. (2009)

Highly scaled up MFC produced, power density of 8 W/m3

Kalathil et al. (2012)

Wang et al. (2008)

Lu et al. (2009)

CercadoQuezada et al. (2010)

Behera et al. (2010)

Kaewkannetra et al. (2011) Velasquez-Orta et al. (2011)

(Continued)

16

Integrated Microbial Fuel Cells for Wastewater Treatment

Table 1.2 (Continued) Type of wastewater

Type of MFCs

Key performances

Food processing

Two chamber

G

Dairy industry

Dual chamber

G

Cheese industry

Tubular

G

G

Distillery

Dual chamber

G

G

Dairy

Dual chamber

G

Combined industrial wastewater: vegetable oil, metal works, and glass and marble industries Surgical cotton industry

Dual chamber

G

G

G

Upflow anaerobic

G

G

G

Dairy

Dual chamber

G

G

G

Agro-food industry

Dual chamber

G

References

Maximum current and power density of 527 mA/m2 and 230 mW/m2, respectively

Mansoorian et al. (2013)

Highest power density of 192 mW/m2, volumetric power of 3.2 W/m3, and COD removal of 91% 80% COD removal from liquid waste, and 60% COD removal from cheese whey sludge Power density of 3.2 6 0.3 W/m3 Power density of 63.8 6 0.65 mW/m2 COD removal of 63.5% 6 1.5% Maximum current density and power density of 3.74 mA and 621.13 mW/m2 COD removal of 85%
90% Maximum voltage of 890 mV Coulombic efficiency of 5184.7 C

Elakkiya and Matheswaran (2013)

TCOD removal of 78.8% and SCOD removal of 69% TSS removal of 62% Maximum power density of 116.03 mW/m2 Maximum voltage of 576 mV Power density of 92.2 mW/ m2 COD removal of 63% 6 1.5% Power density of 27 W/m3

Tamilarasan et al. (2017)

Kelly and He (2014)

Samsudeen et al. (2015)

Mansoorian et al. (2016) Abbasi et al. (2016)

Faria et al. (2017)

Cecconet et al. (2018) (Continued)

Introduction to microbial fuel cells: challenges and opportunities

17

Table 1.2 (Continued) Type of wastewater

Type of MFCs

PRW and LW

Cylindrical

Key performances G

G

Sugarcane molasses

Dual chamber

G

G

G

Fishery industry

Single chamber

G

G

PRW, current and power of 3.35 mA and 1.12 mW LW, current and power of 3.2 mA and 1.02 mW Open circuit voltage of 990 6 5 mV Closed circuit voltage of 453 6 6 mV Power density of 188.5 mW/ m2 Voltage of 750 mV Suspended solid removal of 87%

References Mohanakrishna et al. (2018)

Hassan et al. (2019)

Bancho´n et al. (2019)

FAJ, Fermented apple juice; LW, labanah whey wastewater; PRW, petroleum refinery wastewater; YW, yoghurt waste.

potential use in disinfection applications as well, such as the use of calcium hypochlorite (Momoh and Naeyor, 2010) and sodium hypochlorite (Jadhav et al., 2014) as catholytes in MFC. Arends et al. (2014) explored the disinfection potential of H2O2 in a constructed wetland BES using imposed potential. The reduction of oxygen to hydrogen peroxide in the cathodic chamber and complete disinfection at 0.1% H2O2 concentration within 1 h of contact time was achieved. In a situation where raw water is heavily contaminated with microorganisms, utilization of the cathodic chamber of MFC for disinfection can minimize the cost associated with the treatment processes. There are many pollutants present in wastewater that are recalcitrant in nature. Treatment of such pollutants cannot be addressed in traditional wastewater treatment systems. As a result, there is a need for improvement in current wastewater treatment processes. Advanced oxidation processes such as Fenton oxidation can treat recalcitrant contaminants in wastewaters. Fenton oxidation process employs iron salt and hydrogen peroxide in an acidic medium to produce hydroxyl radicals. Hydroxyl radicals have very high redox potential that has the ability to degrade a wide range of recalcitrant organic compounds present in wastewater. However, Fenton oxidation process is costly due to the need of expensive chemicals such as hydrogen peroxide. To overcome this challenge, MFCs can be used for producing in situ hydrogen peroxide (Asghar et al., 2017). The potential of MFCs for H2O2 production was first explored by Zhu and Ni (2009). Their work reported that a cathode of carbonaceous material could significantly increase the probability of hydrogen peroxide production in the cathode chamber. The production of hydrogen peroxide in MFCs opens up a significant opportunity for developing advanced oxidation process that can be utilized for treating recalcitrant pollutants of wastewater.

18

1.5

Integrated Microbial Fuel Cells for Wastewater Treatment

Conclusion

Until recently, MFCs among all types of BESs have shown some realistic practical promises in terms of wastewater treatment and electricity generation. From here onwards, there are certain areas where future advances should be focused. These are the applicability of technology and the design aspects of the technology in order to meet the criteria of practical real-world application at low cost. There are certain future directions for MFC scale-up. These are to begin with (1) main focus on wastewater treatment as a standalone technology or in conjunction with conventional wastewater treatment technologies; (2) footprint reduction through compacting the system to achieve high power output and to enhance the usability of the system in real-life scenarios; (3) the compactness and reduction in size of units in overcoming ohmic losses and transport limitations; (4) new developments should include MFC power management systems and the incorporation of energy harvesting and storage systems in order to enhance MFC performance for practical use; (5) to ensure the long durability of the system and its components, both internal and external, elements should be resistant to biofouling, scaling and corrosion; and (6) exploration of new applications beyond electricity generation.

Acknowledgment The authors would like to greatly acknowledge the technical support received from Ms. Pratiksha Srivastava from the University of Tasmania.

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7. Abbassi, R., Yadav, A.K., Huang, S., Jaffe, P.R., 2014. Laboratory study of nitrification, denitrification and anammox processes in membrane bioreactors considering periodic aeration. J. Environ. Manage. 142, 53
59. Addi, H., Ramirez, F.M., Martinez, V.M.O., Garcia, M.J.S., Fernandez, F.J.H., Rios, A.P.D. L., et al., 2018. Treatment of mineral oil refinery wastewater in microbial fuel cells using ionic liquid based separators. Appl. Sci. 8, 438. Angelakis, A.N., Snyder, S.A., 2015. Wastewater treatment and reuse: past, present and future. Water 7, 4887
4895. Angosto, J.M., Lopez, J.A.F., Godinez, C., 2015. Brewery and liquid manure wastewaters as potential feedstocks for microbial fuel cells: a performance study. Environ. Technol. 36 (1), 68
78. Ansari, S.A., Parveen, N., Han, T.H., Ansari, M.O., Cho, M.H., 2016. Fibrous polyaniline@manganese oxide nanocomposites as supercapacitor electrode materials

Introduction to microbial fuel cells: challenges and opportunities

19

and cathode catalysts for improved power production in microbial fuel cells. Phys. Chem. Chem. Phys. 18 (13), 9053
9060. Arends, J.B.A., Van Denhouwe, S., Verstraete, W., Boon, N., Rabaey, K., 2014. Enhanced disinfection of wastewater by combining wetland treatment with bioelectrochemical H2O2 production. Bioresour. Technol. 155, 352
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91. Yadav, A.K., Dash, P., Mohanty, A., Abbassi, R., Mishra, B.K., 2012. Performance assessment of innovative constructed wetland-microbial fuel cell for electricity production and dye removal. Ecol. Eng. 47, 126
131. Yang, H., Zhou, M., Liu, M., Yang, W., Gu, T., 2015. Microbial fuel cells for biosensor applications. Biotechnol. Lett. 37, 2357
2364. Yang, N., Zhan, G., Li, D., Wang, X., He, X., Liu, H., 2019. Complete nitrogen removal and electricity production in Thauera-dominated air-cathode single chambered microbial fuel cell. Chem. Eng. J. 356, 506
515. Yazdi, A.A., D’Angelo, L., Omer, N., Windiasti, G., Lu, X., Xu, J., 2016. Carbon nanotube modification of microbial fuel cell electrodes. Biosens. Bioelectron. 85, 536
552. You, S., Zhao, Q., Zhang, J., Jiang, J., Zhao, S., 2006. A microbial fuel cell using permanganate as the cathodic electron acceptor. J. Power Sources 162, 1409
1415.

Introduction to microbial fuel cells: challenges and opportunities

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Yu, B., Tian, J., Feng, L., 2017. Remediation of PAH polluted soils using a soil microbial fuel cell: influence of electrode interval and role of microbial community. J. Hazard. Mater. 336, 110
118. Zhang, F., He, Z., 2015. Scaling up microbial desalination cell system with a post-aerobic process for simultaneous wastewater treatment and seawater desalination. Desalination 360, 28
34. Zhao, S., Li, Y., Yin, H., Liu, Z., Luan, E., Zhao, F., et al., 2015. Three-dimensional graphene/Pt nanoparticle composites as freestanding anode for enhancing performance of microbial fuel cells. Sci. Adv. 1 (10), e1500372. Zhao, H., Zhao, J., Li, F., Li, X., 2016. Performance of denitrifying microbial fuel cell with biocathode over nitrite. Front. Microbiol. 7, 1
7. Zhao, Y., Ma, Y., Li, T., Dong, Z., Wang, Y., 2018. Modification of carbon felt anodes using double-oxidant HNO3/H2O2 for application in microbial fuel cells. RSC Adv. 8, 2059
2064. Zheng, S., Yang, F., Chen, S., Liu, L., Xiong, Q., Yu, T., et al., 2015. Binder-free carbon black/stainless steel mesh composite electrode for high-performance anode in microbial fuel cells. J. Power Sources 284, 252
257. Zhou, M., Chi, M., Luo, J., He, H., Jin, T., 2011. An overview of electrode materials in microbial fuel cells. J. Power Sources 196, 4427
4435. Zhu, X., Ni, J., 2009. Simultaneous processes of electricity generation and p-nitrophenol degradation in a microbial fuel cell. Electrochem. Commun. 11, 274
277. Zhuang, L., Zheng, Y., Zhou, S., Yuan, Y., Yuan, H., Chen, Y., 2012. Scalable microbial fuel cell (MFC) stack for continuous real wastewater treatment. Bioresour. Technol. 106, 82
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777.

Microbial fuel cell integrated wastewater treatment systems

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Pratiksha Srivastava1, Rouzbeh Abbassi2, Asheesh Kumar Yadav3, Vikram Garaniya1 and Faisal Khan4 1 Australian Maritime College, College of Sciences and Engineering, University of Tasmania, Launceston, TAS, Australia, 2School of Engineering, Faculty of Science and Engineering, Macquarie University, Sydney, NSW, Australia, 3Academy of Scientific and Innovative Research (AcSIR), CSIR-Human Resource Development Centre, (CSIR-HRDC) Campus, Ghaziabad, India; Environment and Sustainability Department, CSIR-Institute of Minerals and Materials Technology, Bhubaneswar, India, 4Centre for Risk, Integrity and Safety Engineering, Faculty of Engineering and Applied Science, Memorial University of Newfoundland, St. John’s, NL, Canada

Chapter Outline 2.1 Introduction 29 2.1.1 2.1.2 2.1.3 2.1.4 2.1.5

Sediment microbial fuel cells 30 Constructed wetlands-microbial fuel cells 32 MBR-microbial fuel cells 35 Desalination cell-microbial fuel cells 39 Other processes 40

2.2 Conclusion 41 References 42

2.1

Introduction

Microbial fuel cells (MFCs) work on the principle of anodic oxidation and cathodic reduction. MFCs have gained a lot of attention for wastewater treatment over the past decades; however, most of them are used on a laboratory scale. The upscaling of MFC is not yet finalized as a single unit. Many researchers have used stacked MFCs for achieving the goal of upscaling (Feng et al., 2014). The upscaling/field applicability of MFCs is a challenge due to many reasons. The major shortcoming of MFC is in the cathode region (Zhao et al., 2006), less availability of terminal electron acceptor at the cathode, limits the flow of electrons. Most of the MFCs depend on the dissolved oxygen present at the cathode in the form of terminal electron acceptor. The discharging electrode at the cathode plays a vital role for the transfer of electrons, and many researchers have used a platinum catalyst at the cathode for better performance and to reduce the cathode challenges. However, using Pt is again a costly affair and not suitable to use in MFCs in most cases Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00002-3 © 2020 Elsevier Inc. All rights reserved.

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Integrated Microbial Fuel Cells for Wastewater Treatment

(Logan and Regan, 2006; Zhao et al., 2006). The second challenge is the size of the MFCs. As the size of the MFC increases, the internal resistance of the systems increases as well, which results in lower power density. According to Logan and Regan (2006) when internal resistance of a system decreases, power density of the system increases, which gives more possibility of electricity generation. The other problem is its maintenance on a larger scale, such as electrode fouling, beyond the laboratory (Logan, 2010). However, for expanding applicability of MFC, several integrated wastewater treatment technologies have been introduced. Many other low-cost wastewater treatment technologies such as desalination cells (DS) and membrane biofilters (MBR) were lacking its application due to system design and other shortcomings. These were overcome by integrating MFCs into them. Due to its suitable design compatibility, the MFC has been integrated into several other technologies for the enhancement of treatment efficiency with simultaneous electricity generation. With this concept as a foundation, the anode portion oxidizes organics in anaerobic conditions and cathode acts as a terminal electron acceptor and with the potential difference between both chambers electricity generation occurs (Logan, 2008). The MFC had been integrated in the existing designing of the other technologies. The integrated wastewater treatment technologies that emerged in a recent year are as follows: sediment MFCs (SMFCs), constructed wetlands-MFCs (CW-MFCs), membrane bioreactors-MFCs (MBR-MFCs), DS-MFCs, and others. Furthermore, there are several processes that are also based on the integration of MFCs, that is, hybrid processes such as electro-Fenton reactions and algal bioreactors.

2.1.1 Sediment microbial fuel cells The configuration of MFC allows its integration into many technologies such as SMFCs, which have the same configuration as a normal MFC. In SMFCs, anode can be buried into the sediment (anaerobic) and cathode (aerobic) at the upper part where oxygen is present as a terminal electron acceptor as shown in Fig. 2.1. More interestingly, until 2001, it was not noticed that microbes can donate their electron exogenously. Reimers et al. (2001) explored the possibility of extracellular electron transfer by implementing an MFC into marine sediment (Logan, 2008). They studied exoelectrogenic electron transfer by putting an electrode into the sea sediment to harness energy. The main strategy behind SMFCs was in inserting the anode portion into the sediment (anaerobic) part of the sea and the cathode at the surface of the water. Reimers et al. (2001) concluded that exoelectrogenic bacteria are already present in the sediment and can donate electrons to the electrode in anaerobic conditions. The electrons from the sediment can be transferred to the cathode where dissolved oxygen in the surface water is already present and that can work as a terminal electron acceptor. They achieved power density of 50 mW/m2 from the first SMFCs. The configuration of all types of SMFCs was based on the anaerobic and aerobic portion at the sediment and the surface, respectively. He et al. (2007) explored the possibility of rotating a cathode electrode to provide more oxygen in river sediment and achieved power density of 49 mW/m2 in comparison to a nonrotating cathode, which achieved 29 mW/m2. In this case the rotating cathode was providing more oxygen at the cathode, and that is why the transfer rate of electrons

Microbial fuel cell integrated wastewater treatment systems

31

Figure 2.1 General design of SMFCs integrated with MFCs. MFC, Microbial fuel cell; SMFC, sediment MFC.

increased from the anode. Alternately, in the nonrotating cathode, it was understood that there was limitation of terminal electron acceptors at the cathode. Bond et al. (2002) tested three laboratory scale two-chambered MFCs with marine sediment in anode, and seawater in the cathode, and found an average power production of 0.016 W. They have found a specific type of microbes, Geobacteraceae, which use electrodes as an electron acceptor in anaerobic conditions. These microbes oxidize organics for the support of their growth and use electrodes as an electron acceptor. The electricity generated from SMFCs after storage has been also used for sensor purposes. Donovan et al. (2008) used an SMFC in Palouse River, Pullman, Washington, United States, to power a wireless sensor. The wide availability of organics at the sediment provides a wider application to SMFCs particularly for marine sensors and underwater vehicles. Instead of using a battery that mainly has a short life span, SMFCs supplied a continuous energy supply after storage of energy from SMFCs (Wang and Ren, 2013). With time, different configurations and designs of SMFCs evolve, such as benthic MFCs (BMFCs), floating macrophytes based MFCs (FMFCs), soil-MFCs (SLMFCs) (Xu et al., 2015), etc. Moreover, BMFCs are also an extended application of SMFCs in a marine ecosystem, particularly for bioremediation of pollutants such as sulfur and organic matter at the sediment (Li and Yu, 2015). The pollutants present at the sediment of BMFCs act as a fuel for the microbes, and oxidation of pollutants at the sediment and reduction of oxygen at the surface generates electricity. The continuous recording of the generated voltage and storage of the electricity from BMFCs can also be used for an online monitoring in ocean as reported by Donovan et al. (2013) and Imran et al. (2019). Similar to SMFCs, a few researchers have also tried BMFCs as a power supply for sensor purposes (Nielsen, 2009; Tender et al., 2008).

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Integrated Microbial Fuel Cells for Wastewater Treatment

Many studies have also reported the modification of electrodes for achieving higher electricity generation from the SMFCs and BMFCs, a summary of these studies is given in Table 2.1. Furthermore, the design strategy of FMFCs is inspired by the SMFCs and plant MFCs (PMFCs) design. In PMFCs, plants secrete rhizodeposits (organics and nutrients) to the sediment. The secreted rhizodeposits utilized by the microbes as a fuel at the sediment, and oxygen as a reductant at the surface, allows electricity generation (Schamphelaire et al., 2008; Srivastava et al., 2018a). With the said motivation a sediment FMFCs was introduced by Mohan et al. (2011) for treating high volatile fatty acids (VFA) and residual organic matter from the outlet of an H2 bioreactor. The chemical oxygen demand (COD) and VFA removal were around 86.67% and 72.32%, respectively. The current generation was found to increase with the increase in COD and achieved up to 224.93 mA/m2. The said study was miniaturized and found to be efficient for organics removal and electricity generation. In addition, SL-MFCs are also motivated with the SMFCs for toxic pollutant removal from the waterlogged soil (Huang et al., 2011). In the waterlogged conditions at the sediment, the unavailability of electron acceptors makes pollutants’ removal difficult, and in that condition other electron acceptors such as conductive 2 materials, SO32 4 ; and NO3 , act as an artificial electron acceptor. However, the addi2 tion of extra chemicals in anode portion such as SO32 4 ; and NO3 is not practical, because later these chemicals also need to be removed from the wastewater. The microbes present in the anodic region donate their electron on the conductive materials present in the sediment of soil and the electron travels to the cathode through the electric wire. That is how the pollutant in the logged condition is degraded. With the aim of in situ degradation of pollutant, Huang et al. (2011) evaluated an SL-MFC for enhanced anaerobic phenol degradation. They achieved 90.1% phenol removal with the highest power density of 29.45 mW/m2 in close circuit condition. The removal was also correlated with COD, and it was found that the removal of phenol was due to the MFC integration. Another study conducted an experiment on SL-MFC for the removal of organic pesticide and hexachlorobenzene (HCB). This study found 71.15% removal along with power density of 77.5 mW/m2 (Cao et al., 2015). The study concluded that the presence of electrodes promoted the electrogenic bacteria to provide more electrons, and that is why removal of HCB increased. Deng et al. (2014) studied the factors that mainly affect the power generation of SL-MFCs. According to their findings, the depth of soil for anode positioning and temperature affect the power generation. They concluded that with more depth of soil, ohmic resistance increases, and that is one of the key reasons for low power generation.

2.1.2 Constructed wetlands-microbial fuel cells Constructed wetland (CW) is a low-cost passive treatment technology for the treatment of wastewater (Yadav et al., 2012a). However, it is still not a preferred technology for the treatment of wastewater in many cases. The main reason for low preference is its low treatment efficiency and a large land area requirement for

Microbial fuel cell integrated wastewater treatment systems

33

Table 2.1 Different forms of sediment microbial fuel cells (SMFCs), based on electricity generation and electrode materials. Types of MFCs

Electrodes used

Electricity generation

References

BMFCs

Anode 5 graphite rod, cathode 5 graphite plate associated with carbon fiber/titanium wires

Reimers et al. (2006)

SMFCs

Anode 5 carbon cloth, wrapped with titanium wire, cathode 5 carbon paper

BMFCs

SMFCs

Carbon fiber densely packed in twisted titanium wires Graphite felt

P-SMFCs

Graphite mats

SMFCs

P-SMFCs

Anode 5 carbon cloth, cathode 5 carbon paper/Pt Graphite mat

Normalized power density (P.D.) to the anode 5 34 mW/m2, normalized P.D. to the seafloor 5 1100 mW/m2 Chitin as a substrate, P. D. 5 76 6 25 and 84 6 10 mW/m2, cellulose as a substrate, P.D. 5 83 6 3 mW/m2 Continuous P. D. 5 233 mW/m3, and highest P.D. 5 3.8 W/m3 Current density (C.D.) 5 20.2 6 0.93 mA/m2 Highest voltage 5 102 mV, P.D. 5 1.3 mW/m2 P.D. 5 2162 mW/m3 Voltage 5 500 708 mV

SMFCs

Graphite plate

P.D. 5 10.4 6 0.3 mW

SMFCs

Graphite felt

P.D. 5 2.33 mW

SMFCs

Graphite plate

SMFCs

Anode 5 Graphite plates, cathode 5 Pt-coated graphite felt Anode 5 Porous graphite, cathode 5 graphite felt Graphite rod

Aerobic SMFC, P. D. 5 6.02 6 0.34 and anaerobic SMFC 3.63 6 0.37 mW/m2 P.D. 5 55.2 6 5.9

Chen et al. (2012) Donovan et al. (2013) Ewing et al. (2014) Sherafatmand and Ng (2015)

SMFCs PMFCs

P-SMFCs

Anode 5 carbon fiber felt, cathode 5 graphite disk

Power 5 55 6 2 µW P.D. 5 29.78 mW/m3, C. D. 5 610 mA/m2 Output voltage 5 0.32 V

Rezaei et al. (2007)

Nielsen et al. (2007) Hong et al. (2010) Chen et al. (2012) Morris and Jin (2012)

Zhao et al. (2016) Ewing et al. (2017) Srivastava et al. (2018a) Zhu et al. (2019) (Continued)

34

Integrated Microbial Fuel Cells for Wastewater Treatment

Table 2.1 (Continued) Types of MFCs

Electrodes used

Electricity generation

References

SMFCs

Anode 5 carbon fiber, cathode 5 carbon fiber brush Carbon felt electrodes

P.D. 5 108.89 mW/m2

Wu et al. (2019)

Pt-coated cathode 5 44.88 mW/m2, cerium-coated cathode 5 26.77 mW/m2, cerium- and Pt-coated cathode 5 23.61 mW/m2, cerium-coated anode and cathode 5 20.25 mW/m2, cerium- and Pt-coated cathode 5 17.24 mW/m2, unmodified control 5 10.94 mW/m2

Imran et al. (2019)

BMFCs

construction. To improve the treatment efficiency of CW and decrease the land area requirement, Yadav (2010) developed a hybrid technology that integrates with MFC into CW. In the past 7 8 years, it has gained a lot of attention by researchers across the world. The low treatment efficiency of CW is due to the fact that most of the portion in CW is mainly anaerobic (less electron acceptor) (Srivastava et al., 2018b). Therefore the primary goal of integration of MFC into CW is to enhance the treatment performance of the CW. Lately however, electricity generation was also found to be another resource from CW (Yadav et al., 2018). The anaerobic condition at the bottom and aerobic condition at the surface of CW allow MFC for successful integration as shown in Fig. 2.2. In this case the presence of conductive materials at the bottom acts as an artificial electron acceptor, and the presence of conductive materials at the surface acts as a cathode, which by oxygen reduction generates electricity (Yadav et al., 2012b). The microbes at the bottom of CWMFCs oxidize organics and donate electrons to the electrode, and the electron flows from anode to cathode through an electric wire. With the potential difference between both chambers, electricity generation occurs and so the treatment efficiency increases. The electrode at the bottom mainly acts as an artificial electron acceptor for the microbes in the anode portion. The first batch study on CW-MFC was studied for dye removal in a vertical flow CW by our group. The study reported maximum power density around 69.75 mA/m2, maximum COD of 70% and dye removal of 93.15% (Yadav et al., 2012b). In a subsequent year, Villasenor et al. (2013) studied a horizontal flow CW on different organic loading rates and COD removal of 90% 95% was achieved with maximum power density of 43 mA/m2 and coulombic efficiency (CE) of 0.45%. In the same year, Zhao et al. (2013) studied batch and continuous vertical

Microbial fuel cell integrated wastewater treatment systems

35

Figure 2.2 Integration of MFCs into the CW (CW-MFCs): (A) MFC, (B) CW, and (C) CWMFC. CW-MFC, Constructed wetlands-microbial fuel cell; MFC, microbial fuel cell.

CW for swine wastewater and reported that batch study removed 71.5% of COD and with power density of 12.83 µW/m2. In continuous mode however, removal was 76.5% and power density 9.4 mW/m2. Later, Corbella Clara et al. (2015) examined operational, design, and microbial aspects in CW-MFCs, the study observed 36 mW/m2 of power density and the presence of 13% 16% Geobacter in horizontal subsurface CW implemented with MFC. Moreover, the electricity generation of CW-MFCs is still not high, and CE was found to be very low. The low CE in CW signifies that the treatment efficiency of CW does not just increase by donating electrons and there are some other processes involved in it which need to be further explored as suggested by Srivastava et al. (2018b). Furthermore, the integration of MFC into CW has been studied for different wastewater treatments, different flow regimes, enhancement of electricity production, testing of different materials, and electrode positioning as given in Table 2.2. Recently Ramı´rez-Vargas et al. (2018) and Doherty et al. (2015a) reviewed integration of MFC into CW from its evolution and development. Still the development of CW-MFCs is in progress, and many areas still need to be explored before its larger application. However, a project named METlands is already running on a field scale with the concept of integration of conductive material into the wetlands (Salas et al., 2017).

2.1.3 MBR-microbial fuel cells In recent years, MBR has gained a lot of attention for wastewater treatment and has proven to be an efficient technology (Abbassi et al., 2014). The two key points of MBR systems are activated sludge process and membrane separation. The membrane in MBR acts as a separation technique followed by the biological processes (activated sludge). The major advantage of MBR is that it has a smaller footprint with less sludge generation. Despite achieving higher treatment efficiency, MBR faces the problem of membrane fouling. Logan (2008) introduced the idea of integrating MFCs into MBR to overcome the shortcomings of MBR for wastewater

Table 2.2 Microbial fuel cell (MFC) integrated constructed wetlands (CW): flow regime, electricity generation, design configuration, electrode materials. Types of CWMFCs

Anode and cathode specifications

Initial carbon concentration

Removal

Electricity generation

Reference

Vertical FlowMFCs, batch mode

Anode surface area 5 40.93 cm2, cathode surface area 5 40.93 cm2

COD removal 5 up to 75%, dye removal 5 80.0% 93.15%

Power density (P.D.) 5 15.73 mW/m2 and C.D. 5 69.75 mA/m2

Yadav et al. (2012b)

Horizontal subsurface flow, continuous mode Horizontal subsurface flow, continuous mode Upflow CW

Stainless steel anode 5 7.5 cm2 and cathode 5 7.5 cm2

8 g/L sucrose, methylene blue dye concentration 5 500, 1000, 1500, and 2000 mg/L Biological oxygen demand (BOD) 5 115 mg/L, COD 5 235%, ammonia 5 39% COD 5 250 1100 mg/L

Ammonia 5 60%, COD 5 71%, and BOD 5 54%

P.D. 5 36 mW/m2 and C.D. 5 21 mA/m2

Corbella Clara et al. (2015)

COD removal 5 80% 95%

P.D. 5 43 mW/m2, C.D. 5 37.1 mA/m2, and CE 5 0.45%

Villasenor et al. (2013)

Upflow downflow CW Upflow vertical CW

Vertical CW

Graphite anode and graphite cathode

Carbon felt

COD 5 314.8 6 13 mg/L

COD removal 5 100%

P.D. 5 6.12 mW/m2

Graphite

COD 5 411 854 mg/L

COD removal 5 81%

P.D. 5 0.268 mW/m3

Anode 5 graphite and charcoal, cathode 5 Pt catalyst containing carbon cloth Anode 5 graphite felt, cathode 5 Pt-coated carbon cloth

Glucose 5 250, 500, and 750 mg/L

COD removal on 750 mg/L 5 86%

P.D. 5 320.8 mW/m3 and C.D. 5 422.2 mA/m3

Oon et al. (2015) Doherty et al. (2015b) Srivastava et al. (2015)

Glucose 5 700 and 2000 mg/L

COD removal 5 72.17% and 46.77%

C.D. 5 181.38 mA/m3 and P.D. 5 31.04 mW/m3

Srivastava et al. (2017)

Vertical upflow

Anode 5 stainless steel, cathode 5 granular activated carbon

Glucose 5 50 mg/L and methyl orange 5 450 mg/L

Vertical CW

Anode 5 cylinder of titanium mesh filled with activated carbon, cathode 5 titanium mesh

Sucrose 5 53.483 mg/L, (NH4)2SO4 5 37.714 mg/L, and KNO3 5 50.5 mg/L

Upflow vertical CW

Calcined petroleum coke

COD inlet loads ranging from 8.9 to 142 g/ (m2 day)

Horizontal flow bed

Anode 5 stainless steel mesh, cathode 5 carbon felt

Organic loading rate (OLR) 5 4.9 6 1.6, 6.7 6 1.4, and 13.6 6 3.2 g COD/ (m2 day)

Decolorization rate 5 69.41%, COD removal rate 5 15.62% Total Nitrogen 5 82.46% 6 4.74%, COD 5 82.32% 6 12.85%, NH41 N 5 77.79% 6 10.26% and Total Phosphorus 5 95.06% 6 5.45% BOD5 5 88%, COD 5 90%, NH4 N 5 46%, and PO4 P 5 86% Increased COD 5 60% 70%, ammonium 5 25% 40%

Highest P.D. 5 0.688 W/m3

Fang et al. (2016)

Average voltage 5 265.77 6 12.66 mV, P.D. 5 3714.08 mW/m2, C.D. 5 16.63 mA/m2

Xu et al. (2018)

Average current 5 0.126 mA/cm

Ramı´rez-Vargas et al. (2019)

Low OLR 5 33 6 6 mA/ m2, high OLR 5 43 6 10

Hartl et al. (2019)

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Integrated Microbial Fuel Cells for Wastewater Treatment

Figure 2.3 General design of an MFC-integrated MBR reactor. MFC, Microbial fuel cell.

treatment and electricity generation for at the first time. It suggested that MFC can be combined with the MBR system as a posttreatment process. As shown in Fig. 2.3, the organics present in the wastewater act as feed for microbes at the anode and the presence of oxygen at the cathode for reduction of oxygen. The membrane fouling in MBR mainly occurred due to the interaction of pollutant with the membrane materials. To overcome the problem of membrane fouling, Wang et al. (2016) integrated the MFC system into the MBR. In comparison to the conventional MBR systems, the fouling of membrane in MFC-integrated MBR was significantly reduced. Initially Cha et al. (2010) inserted MFC into the activated sludge process and considered an aerated tank as a biocathode for MFC. The achieved highest CE from the study was 39.6%, with a higher power density of 16.7 W/m3. The generation of voltage was very much influenced from the aeration in the system, even at a very low concentration of oxygen the voltage rose from 30 to 200 mV. Later, Wang et al. (2012) also proved the role of aeration in MFC integrated into MBR reactor, where the aerated tank acted as a cathode of MFC and achieved 1.9 6 0.4 mA of average current, with a COD removal of 89.6% 6 3.7%. The study suggests that the integrated systems use better utilization of oxygen and increased the treatment efficiency of the wastewater. In addition, Su et al. (2013) combined sludge MFC into the MBR system particularly for sludge reduction, wastewater treatment, membrane fouling, and energy recovery. The study achieved 5.1% higher sludge reduction and more than 90% removal for COD and ammonia than the traditional MBR system. Finally, the study concluded that the integration of MFC into MBR increased the treatment efficiency along with electricity generation as well as reduced membrane fouling problems. Malaeb et al. (2013) used a hybrid air-cathode MBR-MFCs for the wastewater treatment and simultaneous ultrafiltration. The maximum power density with air-cathode was 0.38 W/m2 with 97% removal of COD and 97% NH3 N along with 91% total bacteria. Their findings illustrate that even with less aeration or less energy input to the MBR-MFCs,

Microbial fuel cell integrated wastewater treatment systems

39

the treatment of wastewater can be high with simultaneous electricity generation. Tan et al. (2017) mentioned that in anaerobic MBR systems temperature and integration of MFC have a significant effect on MBR. According to their findings the maximum COD removal efficiency was achieved in a mesophilic condition at a temperature of 45 C. They also reported that at the same temperature (45 C), the system has more resistivity for filtration, and higher organic removal can be achieved.

2.1.4 Desalination cell-microbial fuel cells Desalination (DS) is another alternative for the treatment of wastewater and to produce potable water. The processes involved in DS are distillation, electrodialysis, and reverse osmosis. Apparently, due to the high-energy requirement and maintenance, it is not applicable in all situations. In the past few decades, with continuous improvement in the technology, it is now more reliable and efficient as a treatment. However, energy requirement for the whole operation is still a challenge (Cao et al., 2009). The advantages of MFCs for the generation of renewable energy such as methane, hydrogen, and electricity production give it the possibility for implementation of MFC into DS. Cao et al. (2009) introduced a new type of DS, called microbial DS, which is based on transfer of ions from water with a proportion to the electron transferred by the bacteria. As shown in Fig. 2.4, the three-chambered DS-MFCs were designed with an anion exchange membrane and cation exchange membrane along with a middle chamber. The microbes degrade pollutants and generate electricity in the anode chamber, and negatively charged ions move from the middle chamber to the anode through a membrane. Similar to normal MFCs, the cathode consumes protons and a positive charge from the middle chamber moved to the cathode. Because of this, water in the middle chamber is desalinated with

Figure 2.4 The general design of DS-MFCs. DS-MFC, Desalination cell-microbial fuel cell.

40

Integrated Microbial Fuel Cells for Wastewater Treatment

simultaneous wastewater treatment. Mohanakrishna et al. (2010) used DS-MFCs for the treatment of distillery wastewater, along with current production of 2.12 2.48 mA, and it was efficient enough to treat 72.84% of COD and color reduction of 31.67%. Zhang and He (2013) designed an osmotic MFC with a forward-osmosis membrane integrated with microbial desalination cell (MDC) for artificial wastewater and saline water treatment. The study achieved power production of 0.160 kW h/m3 along with conductivity reduction of 95.9%. The system proved to be an efficient system for salt removal due to dilution and desalination. Sevda and Abu-Reesh (2019) implemented osmotic MFC with upflow MDC for seawater treatment and used organics for MFC from petroleum wastewater. Their study found that the combined system can remove up to 93% COD with 48% salt from seawater. However, there are still challenges for the scaling up of DS-MFCs or MDC. Zhang and He (2015) tried a scaling up of the system using 105 L MDC. The finding from the study is that the multiple feeding inlets could enhance the current generation. By applying external voltage, the current generation increased from 670 to 2000 mA and salt reduction of 3.7 9.2 kg/(m3 day). Furthermore, energy requirement for the salt reduction decreased when applied voltage increased; however, for the wastewater treatment energy was still required.

2.1.5 Other processes Many studies have integrated MFCs in a stacked manner for upscaling, and for achieving efficient treatment efficiency. Feng et al. (2014) implemented a series of MFCs together in a stacked manner in a plug flow condition. The flow direction was horizontal with a total volume of 250 L. The maximum current achieved was 0.435 6 0.010 A in each module while treating domestic wastewater with a COD and total nitrogen (TN) removal of 79% 6 7% and 71% 6 8%, respectively. In addition, there are several other processes such as electro-Fenton reactions, in which MFC has been integrated for the enhancement of treatment performance. The strategy behind integration of MFCs to enhance the electro-Fenton reactions is to supplement hydroxyl ions to the Fenton reactions for making reactions faster. In MFC, at the cathode, formation of H2O2 has been reported in many studies, the formation of H2O2 at the cathode can be possible by two or four electron transfer mechanisms (Dong et al., 2018), whereas in electro-Fenton reactions ferrous ions oxidize to ferric ions with the help of oxidants. In MFC-integrated systems the continuous generation of H2O2 in the presence of iron forms a strong oxidant, hydroxyl ions. Therefore at the cathode of MFC in the presence of high oxidant, ferrous ions can be oxidized to ferric ions and release hydroxyl ions oxidize pollutants (Dios et al., 2014). Dios et al. (2014) used BMFCs for the degradation of dye with electro-Fenton reactions and achieved 92.2% of decolorization at the voltage of 1168 mV. Their study concluded that with electroFenton reactions, dye decolorization along with voltage generation increased significantly. Wang et al. (2017) used MFC-Fenton system for the treatment of four emerging contaminants, namely, bisphenol A, estrone, sulfamethazine, and triclocarban. Their study experimented in batches and continuous modes and found that

Microbial fuel cell integrated wastewater treatment systems

41

Figure 2.5 Typical diagram of algae-assisted cathode in MFCs. MFC, Microbial fuel cell.

in both modes H2O2 production initially increased, but after a certain time it decreased. The study concluded that maintaining pH at 3 promotes the acclimatization of H2O2, subsequently the removal. The more oxidant that is formed at the cathode, the more efficient is the reaction rate of degradation. Similarly, algal bioreactors are also implemented with MFCs and the bases of MFC-integrated algal bioreactors are based on solar light; thus the presence of algal biomass converts sunlight into chemical energy. Further the available algae use the chemical energy to again convert into electrical energy. However, the major goal was the enhancement of electricity generation instead of wastewater treatment (Xu et al., 2015). The integrated MFCs into algal bioreactors are called photosynthetic fuel cells or solar cells (De Schamphelaire and Verstraete, 2009; Strik et al., 2008; Xu et al., 2015). In the other configuration of algal MFCs, microbes oxidize the organics at the anode, and at the cathode algae releases oxygen to fulfill the need of aeration (refer to Fig. 2.5). Therefore it is considered as a passive aeration from algae, which supplies oxidant to the cathode for the completion of reactions at the anode. A study has been conducted by del Campo et al. (2013) on the aforementioned concept, with Chlorella vulgaris (algae) at the cathode. The highest achieved power density was 13.5 mW/m2; however, the study concluded that resistance of the cathode was higher than the anode (del Campo et al., 2013). Other researchers have also demonstrated the concept while using blue-green algae used at the cathode for a chemical oxidant. The highest current and power density achieved was 149.5 and 78.12 mW/m2, respectively (Yadav et al., 2015). Other studies demonstrate the feasibility of integration of MFCs into a rice paddy field for the generation of electricity while using algae as an oxidant. The study considered a self-sustainable system and power density and achieved current of 29.78 mW/m3 and 610 mA/m2, respectively, has been reported without using any artificial aeration at the cathode (Srivastava et al., 2018a).

2.2

Conclusion

The integration of MFCs into other technologies has, to a certain extent, been proven successful for the enhanced treatment efficiency. However, the amount of

42

Integrated Microbial Fuel Cells for Wastewater Treatment

electricity generation is still a challenge, as the generation of electricity is mostly dependent on the terminal electron acceptor such as oxygen at the cathode, surface area of the electrode, and internal and external resistance (Xu et al., 2015). In correlation with the pollutant removal the achieved power and current density from integrated systems is still low. The other shortcomings of low electricity generation are found to be dependent on electrode materials, reactor configuration, electrode positioning, and spacing. The possibilities of using the generated electricity from the hybrid system such as SMFCs are after storing the power, which can be useful for underwater vehicles or for sensing purposes. The design of integrated MFC is still a challenge for scalability due to several shortcomings such as when the reactor size increases, resistance of the reactor also increases which mainly hinders the electricity generation. The distance between the electrodes also has an effect on electricity generation (Cheng et al., 2006). Further, the hybrid technologies are also capable of producing value-added products from the waste such as oxidant, self-buffering capacity, which is useful for wastewater industries. For example, production of oxidants such as H2O2 at cathode is useful for enhancing the electro-Fenton reactions. Similarly, incorporation in DS improves the reduction of salinity from seawater. However, the challenges are still not yet solved to bring the technology to field level (Santoro et al., 2017). The challenges of electricity generation with the removal rate as well as configuration and design of the reactor still need to be addressed before it can be used with real field application.

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del Campo, A.G., Can˜izares, P., Rodrigo, M.A., Ferna´ndez, F.J., Lobato, J., 2013. Microbial fuel cell with an algae-assisted cathode: a preliminary assessment. J. Power Sources 242, 638 645. Deng, H., Wu, Y.-C., Zhang, F., Huang, Z.-C., Chen, Z., Xu, H.-J., et al., 2014. Factors affecting the performance of single-chamber soil microbial fuel cells for power generation. Pedosphere 24 (3), 330 338. De Schamphelaire, L., Verstraete, W., 2009. Revival of the biological sunlight-to-biogas energy conversion system. Biotechnol. Bioeng. 103 (2), 296 304. Dios, M.A.F.D., Iglesias, O., Bocos, E., Pazos, M., Sanroma´n, M.A., 2014. Application of benthonic microbial fuel cells and electro-Fenton process to dye decolourisation. J. Ind. Eng. Chem. 20 (5), 3754 3760. Doherty, L., Zhao, Y., Zhao, X., Hu, Y., Hao, X., Xu, L., et al., 2015a. A review of a recently emerged technology: constructed wetland microbial fuel cells. Water Res. 85, 38 45. Doherty, L., Zhao, Y., Zhao, X., Wang, W., 2015b. Nutrient and organics removal from swine slurry with simultaneous electricity generation in an alum sludge-based constructed wetland incorporating microbial fuel cell technology. Chem. Eng. J. 266, 74 81. Dong, H., Liu, X., Xu, T., Wang, Q., Chen, X., Chen, S., et al., 2018. Hydrogen peroxide generation in microbial fuel cells using graphene-based air-cathodes. Bioresour. Technol. 247, 684 689. Donovan, C., Dewan, A., Heo, D., Beyenal, H., 2008. Batteryless, wireless sensor powered by a sediment microbial fuel cell. Environ. Sci. Technol. 42 (22), 8591 8596. Donovan, C., Dewan, A., Heo, D., Lewandowski, Z., Beyenal, H., 2013. Sediment microbial fuel cell powering a submersible ultrasonic receiver: new approach to remote monitoring. J. Power Sources 233, 79 85. Ewing, T., Ha, P.T., Babauta, J.T., Tang, N.T., Heo, D., Beyenal, H., 2014. Scale-up of sediment microbial fuel cells. J. Power Sources 272, 311 319. Ewing, T., Ha, P.T., Beyenal, H., 2017. Evaluation of long-term performance of sediment microbial fuel cells and the role of natural resources. Appl. Energy 192, 490 497. Fang, Z., Song, H., Yu, R., Li, X., 2016. A microbial fuel cell-coupled constructed wetland promotes degradation of azo dye decolorization products. Ecol. Eng. 94, 455 463. Feng, Y., He, W., Liu, J., Wang, X., Qu, Y., Ren, N., 2014. A horizontal plug flow and stackable pilot microbial fuel cell for municipal wastewater treatment. Bioresour. Technol. 156, 132 138. Hartl, M., Bedoya-Rios, D.F., Fernandez-Gatell, M., Rousseau, D.P.L., Du Laing, G., Garfi, M., et al., 2019. Contaminants removal and bacterial activity enhancement along the flow path of constructed wetland microbial fuel cells. Sci. Total Environ. 652, 1195 1208. He, Z., Shao, H., Angenent, L.T., 2007. Increased power production from a sediment microbial fuel cell with a rotating cathode. Biosens. Bioelectron. 22 (12), 3252 3255. Hong, S.W., Kim, H.S., Chung, T.H., 2010. Alteration of sediment organic matter in sediment microbial fuel cells. Environ. Pollut. 158 (1), 185 191. Huang, D.-Y., Zhou, S.-G., Chen, Q., Zhao, B., Yuan, Y., Zhuang, L., 2011. Enhanced anaerobic degradation of organic pollutants in a soil microbial fuel cell. Chem. Eng. J. 172 (2 3), 647 653. Imran, M., Prakash, O., Pushkar, P., Mungray, A., Kailasa, S.K., Chongdar, S., et al., 2019. Performance enhancement of benthic microbial fuel cell by cerium coated electrodes. Electrochim. Acta 295, 58 66. Li, W.W., Yu, H.Q., 2015. Stimulating sediment bioremediation with benthic microbial fuel cells. Biotechnol. Adv. 33 (1), 1 12.

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Sherafatmand, M., Ng, H.Y., 2015. Using sediment microbial fuel cells (SMFCs) for bioremediation of polycyclic aromatic hydrocarbons (PAHs). Bioresour. Technol. 195, 122 130. Srivastava, P., Yadav, A.K., Mishra, B.K., 2015. The effects of microbial fuel cell integration into constructed wetland on the performance of constructed wetland. Bioresour. Technol. 195, 223 230. Srivastava, P., Dwivedi, S., Kumar, N., Abbassi, R., Garaniya, V., Yadav, A.K., 2017. Performance assessment of aeration and radial oxygen loss assisted cathode based integrated constructed wetland-microbial fuel cell systems. Bioresour. Technol. 244 (Pt 1), 1178 1182. Srivastava, P., Gupta, S., Garaniya, V., Abbassi, R., Yadav, A.K., 2018a. Up to 399 mV bioelectricity generated by a rice paddy-planted microbial fuel cell assisted with a bluegreen algal cathode. Environ. Chem. Lett. 17, 1045 1051. Srivastava, P., Yadav, A.K., Garaniya, V., Abbassi, R., 2018b. Constructed wetland coupled microbial fuel cell technology: development and potential applications. Microbial Electrochemical Technology. Elsevier, pp. 1021 1036. Strik, D.P., Terlouw, H., Hamelers, H.V., Buisman, C.J., 2008. Renewable sustainable biocatalyzed electricity production in a photosynthetic algal microbial fuel cell (PAMFC). Appl. Microbiol. Biotechnol. 81 (4), 659 668. Su, X., Tian, Y., Sun, Z., Lu, Y., Li, Z., 2013. Performance of a combined system of microbial fuel cell and membrane bioreactor: wastewater treatment, sludge reduction, energy recovery and membrane fouling. Biosens. Bioelectron. 49, 92 98. Tan, S.P., Kong, H.F., Bashir, M.J.K., Lo, P.K., Ho, C.D., Ng, C.A., 2017. Treatment of palm oil mill effluent using combination system of microbial fuel cell and anaerobic membrane bioreactor. Bioresour. Technol. 245 (Pt A), 916 924. Tender, L.M., Gray, S.A., Groveman, E., Lowy, D.A., Kauffman, P., Melhado, J., et al., 2008. The first demonstration of a microbial fuel cell as a viable power supply: powering a meteorological buoy. J. Power Sources 179 (2), 571 575. Villasenor, J., Capilla, P., Rodrigo, M.A., Canizares, P., Fernandez, F.J., 2013. Operation of a horizontal subsurface flow constructed wetland microbial fuel cell treating wastewater under different organic loading rates. Water Res. 47 (17), 6731 6738. Wang, H., Ren, Z.J., 2013. A comprehensive review of microbial electrochemical systems as a platform technology. Biotechnol. Adv. 31 (8), 1796 1807. Wang, Y.-P., Liu, X.-W., Li, W.-W., Li, F., Wang, Y.-K., Sheng, G.-P., et al., 2012. A microbial fuel cell membrane bioreactor integrated system for cost-effective wastewater treatment. Appl. Energy 98, 230 235. Wang, J., Bi, F., Ngo, H.H., Guo, W., Jia, H., Zhang, H., et al., 2016. Evaluation of energydistribution of a hybrid microbial fuel cell-membrane bioreactor (MFC-MBR) for costeffective wastewater treatment. Bioresour. Technol. 200, 420 425. Wang, Y., Feng, C., Li, Y., Gao, J., Yu, C.-P., 2017. Enhancement of emerging contaminants removal using Fenton reaction driven by H2O2-producing microbial fuel cells. Chem. Eng. J. 307, 679 686. Wu, M., Xu, X., Lu, K., Li, X., 2019. Effects of the presence of nanoscale zero-valent iron on the degradation of polychlorinated biphenyls and total organic carbon by sediment microbial fuel cell. Sci. Total Environ 656, 39 44. Xu, L., Zhao, Y., Doherty, L., Hu, Y., Hao, X., 2015. The integrated processes for wastewater treatment based on the principle of microbial fuel cells: a review. Crit. Rev. Environ. Sci. Technol. 46 (1), 60 91. Xu, F., Cao, F.-Q., Kong, Q., Zhou, L.-L., Yuan, Q., Zhu, Y.-J., et al., 2018. Electricity production and evolution of microbial community in the constructed wetland-microbial fuel cell. Chem. Eng. J. 339, 479 486.

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Removal of heavy metals using bioelectrochemical systems

3

Sukrampal, Rohit Kumar and Sunil A. Patil Department of Earth and Environmental Sciences, Indian Institute of Science Education and Research (IISER Mohali), Mohali, Punjab, India

Chapter Outline 3.1 Introduction 49 3.2 Bioelectrochemical systems for heavy metal removal 56 3.2.1 Concept and principle 56 3.2.2 Reduction of heavy metals at the cathode of bioelectrochemical systems 57

3.3 Electrode materials used for heavy metal removal in bioelectrochemical systems 58 3.4 Conventional technologies versus bioelectrochemical systems-based technology for the removal of heavy metals 63 3.5 Conclusion 66 References 67 Further reading 71

3.1

Introduction

Heavy metals are crucial constituents of a vast number of products, crafts, industrial processes, various modern technologies, and many essential biological molecules (Dominguez-Benetton et al., 2018; Nancharaiah et al., 2016). Though they are essential for various activities, their overextraction and overuse through anthropogenic activities (Watts, 2003) along with a partial influence of the natural calamities have resulted in contamination of different environments. Among the various emerging contaminants, heavy metals are the most hazardous ones. With a density greater than 4000 kg/m3, these are a group of elements (includes transition metals, metalloids, lanthanides, and actinides) that exhibit metallic properties. Heavy metals such as Pb, Cd, Cr, Co, As, Zn, Hg, and Ni are the major contaminants that are listed in the Environmental Protection Agency’s list of priority pollutants (Dominguez-Benetton et al., 2018). Those with the density greater than 4000 kg/m3 (Jobin and Namour, 2017) and specific gravity (sg) of five times or more than the water (1 at 4
C/39
F) such as arsenic (5.7 sg), cadmium (8.65 sg), lead (11.34 sg), and mercury (13.596 sg) are considered to be the most hazardous ones (Dash and Das, 2012; Srivastava and Majumder, 2008). Heavy metal contamination of soils is one of the most critical environmental problems as metals may filter through the different subsurface layers and may reach Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00003-5 © 2020 Elsevier Inc. All rights reserved.

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Integrated Microbial Fuel Cells for Wastewater Treatment

the groundwater (Galanis et al., 2009; Perfus-Barbeoch et al., 2002). They may also enter the food chains, which potentially results in their biomagnification in the consumers and can get transformed into persisting compounds with xenobiotic properties. An overview of the major heavy metal contaminants along with their source or origin, route of exposure, contamination level, and associated health effects is presented in Table 3.1. Some of the heavy metals such as Pb, As, Hg, Cr, and Cd show widespread lethal effects even at a lower concentration ranging from µg to mg/L (Wang and Ren, 2014; Wang et al., 2011). For instance, even a few µg/L concentration of methylmercury is lethal to humans (Dash and Das, 2012). At high concentration, they also affect the population of microorganisms in the soil that can, in turn, have adverse and direct effects on soil fertility (Minnikova et al., 2016; Guo et al., 2010; Wu et al., 2012; Singh and Yakhmi, 2014). Excess concentration of heavy metals may also lead to neurological depositions and damage the various important systems such as nervous, transportation, digestion, and immune systems and may also cause chronic inflammatory diseases that in turn triggers the initiation activity of cancer in humans. Furthermore, they may also lead to premature aging and many other characteristic features of a wide range of diseases (Chowdhury et al., 1987; Richards et al., 2007). Most heavy metals also have the potential to persist in environments for longer periods. For instance, Pd persistency in the soil for more than 150 years has been recorded. It may extend to a range of 1505000 years (Nanda, 1995). The half-life of heavy metals can also be longer. For example, more than 18 years of a half-life has been reported for Cd (Forstner, 1995). The persistence and half-life of heavy metals for longer periods aggravate their toxicityrelated issues further. In order to address the environmental and health-related concerns of heavy metals, various types of physical, chemical, electrochemical, and biological methods have been tested and are under investigation for their removal from the contaminated sites (Table 3.2). These include, for example, ion-exchange, chemical precipitation, electrolysis, electrodialysis, membrane filtration, photocatalysis, and electrochemical treatment technologies (Pedersen et al., 2003; Fu and Wang, 2011). The bio-based approaches have also been extensively researched and proposed for cleaning up of heavy metalcontaminated environments (Khan and Camille, 2017). These are based on the utilization of different kinds of microorganisms and plants and are referred to as microbial remediation and phytoremediation technologies, respectively. All these technologies have several disadvantages and limitations. These include, for example, high cost, environmentally destructive, timeconsuming, technical implication, operational complexities, restricted use, moderate efficiency, foul-smelling, pore chocking, and xenobiotics productionrelated or type problems. A unique approach based on the integration of electrochemistry and microbiology has recently emerged. It is commonly referred to as electrode-assisted bioremediation or electro-bioremediation or bioelectrochemical remediation of heavy metalcontaminated environments. The reactor systems used to achieve this process are referred to as bioelectrochemical systems (BESs). BESs combine microbial and electrochemical processes to convert the chemical energy stored in

Table 3.1 Major heavy metal pollutants—sources, routes of exposure, levels of contamination, and associated health effects. S. no.

Metals

Source of exposure

Major route of exposure

Environmental level

Health effects

Reference

1

Arsenic

Food, air, drinking water or groundwater

Oral

Lung cancer, cardiovascular effects, and encephalopathy

Maher and Butler (1988)

2

Cadmium

Food, cigarette smoking, drinking water, and air

Oral, inhalation and dermal

Glomerular damage, bone mineralization, emphysema

Hutton (1987)

3

Lead

Contaminated food, drinking water, lead-based paint

Inhalation

Mercury

5

Barium

Water, air, dental amalgam fillings, and waste incinerators Wastewater, food, paints, barium barite, and sulfates

Inhalation, dermal, and oral Oral and inhalation

Elevated BP, insomnia, damage CNS, neuropathy, reduced fertility Diarrhea and/or abdominal pain, kidney damage, and acrodynia Abdominal pain, mental illness, balance problems

Hernberg (2000)

4

Air: rural (13 ng/ m3), urban (20100 ng/m3) Drinking water: 0.002 ppm Soil: 25 mg/kg Air: rural (0.15 ng/ m3), urban (215 ng/m3) Drinking water: 5 µg/L, soil: 0.061.1 mg/kg Air: 0.05 µg/m3 water: 510 µg/L soil: 1030 g/kg Air: 0.91.5 ng/m3 water: 0.5100 ng/L soil: 617 mg/kg Water: 0.050 ppm Soil: 0.50 g/kg

6

Silver

Mining, dust, dermal, contaminated food

Oral, dermal, and inhalation

Drinking water: 1 ppb Soil: 0.015 mg/kg

Argyria, silver disease, RussellSilver syndrome

Dash and Das (2012) Jacobs et al. (2002) Gaiser et al. (2009) (Continued)

Table 3.1 (Continued) S. no.

Metals

Source of exposure

Major route of exposure

Environmental level

Health effects

Reference

7

Chromium

Mining, dust, industrial contaminated air, and food

Inhalation, oral, and dermal

Water: 0.05 ppm

Pulmonary sensitization, cancer, sinus, cytotoxicity, and GI problems

8

Uranium

Nuclear exposure, army waste, mining

Inhalation, dermal

Cancer, genotoxic, allergy, kidney damage, digestion, and GI problems

9

Nickel

Inhalation, oral

10

Vanadium

Industrial waste, accidental exposure, drinking water, contaminated food Burning of fossil fuels, automobile waste, mining

Air: insoluble (0.25 mg/m3), soluble (0.05 mg/m3) Water: 30 µg/L Normally nontoxic but at higher conc. is toxic Air: 0.05 ng/m3 At workplace 0.1 mg/m3

Mohanty and Patra (2012) Brugge and Buchner (2011)

Inhalation, oral and dermal

Cancer, GI distress, pulmonary fibrosis, skin dermatitis Apoptosis, cytotoxic, cancer

Borba et al. (2006) Mukherjee et al. (2004)

Table 3.2 An overview of the major conventional and bioelectrochemical approaches used for the removal of heavy metals. Technique 1. Physical Soil replacement/ isolation

Process involved

Advantages

Limitations

Applicability

Acceptance

Multimetal sites

Replacing of contaminated soil with noncontaminated soil

1. Potentially remediate highly contaminated soil 2. Can isolate and remove heavy metals 1. Easily applicable 2. Can be used for a different variety of metals 1. Easily applicable 2. Economically effective 3. Environmentalfriendly

Have a negative effect on the environment, costly and produces hazardous waste

Applicable at small scale for long term

Very low, also need a very short time

Effective

Costly, require energy input

Small scale for long term

Very low, also need very short time

Effective

pH-sensitive can work potentially at low permeability

Small scale for long term

Very low, also need very short time

Effective

Vitrification

Producing vitreous materials by reducing metal bioavailability at high temperature

Electrokinetic remediation

By using DC-voltage to remove heavy metals via electrophoresis, electromigration

(Continued)

Table 3.2 (Continued) Technique 2. Chemical Immobilization

Soil washing

3. Biological Phytovolatilization

Phytostabilization

Process involved

Advantages

Limitations

Applicability

Acceptance

Multimetal sites

Producing stable and immobile complexes by adsorption leads to reduce metal mobility and bioavailability

The temporary solution needs permanent monitoring

Up to medium scale but with short term

Highly acceptable, also, need a very short time

Effective but depends on type or site of metal and immobilization amendment

Removal by extractants, stable and mobile complexes

1. Fast and easy to apply 2. Low cost 3. Applicable on a broad spectrum of inorganic pollutants 1. Cost effective 2. Reduce longterm liability

May lead to environmental issues, sometimes effectivity varies

Small scale

Moderately acceptable, also need very short time

Effective

Transfer to atmosphere via plant uptake and utilization

1. Economically effective 2. Less disruptive

Small scale

1. Economically effective 2. Less disruptive

Low-tomoderate acceptance, timeconsuming Low-tomoderate acceptance, require long timing

No

Sequestration in plant roots to reduce metal bioavailability and mobility

Only applicable for volatile metals, may result in other issues Temporary solution, sometimes effectivity varies

Small scale

Very low

Microbial assisted remediation

Use of Microbes to enhance the capacity of plants

1. Requires short time 2. Enhance plant growth and metal uptake

Depends on microbes, soils, plant, and metal types

Large scale

Very high, Require long timing but lesser than other biological methods

Low but more than other biological methods

Transfer electron to the oxidized metal ion and reduces it to the lesser or nontoxic form

1. Economically effective 2. Applicable on a broad spectrum of inorganic pollutants

Different kinds of electrochemical losses are there, only applicable in some places

Small and lab scale

Low-tomoderate acceptance, Require long timing

Effective but depends on type or site of metal, electric potential, and reactor design

4. Bioelectrochemical

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Integrated Microbial Fuel Cells for Wastewater Treatment

biodegradable organic matter into electricity, hydrogen, or other useful chemicals or to drive processes such as saltwater desalination, and the removal of various contaminants.

3.2

Bioelectrochemical systems for heavy metal removal

Electrochemical treatment processes work on the principle of electrolysis in which, by providing energy from an applied source, metal oxidation at the anode and metal reduction at the cathode are facilitated at specific electric potentials. Though this technique is effective, it requires a lot of maintenance, care, and high energy input and thus involves substantial operational costs. To overcome some of the disadvantages of this approach, researchers started exploring BESs, which utilize microorganisms to mediate, facilitate, or catalyze the redox reactions at both or either of the electrodes (Dermentzis et al., 2011). Several studies have documented that the microorganisms are effective biocatalysts to influence the mobility of metal ions in natural and engineered or designed systems under controlled conditions (Maas et al., 2005; Francis and Nancharaiah, 2015; Nancharaiah and Lens, 2015). BESs open up an opportunity for the development of a novel and useful biotechnology based on the use of microorganisms in electrochemical systems for the removal and recovery of heavy metals from the contaminated soil, sediment, and water environments (Wang and Ren, 2014).

3.2.1 Concept and principle BESs consist of the anode and cathode electrodes that are usually separated by an ion-selective membrane. Microorganisms that possess extracellular electron transfer capabilities are generally used to catalyze either the substrate oxidation or reduction reactions at the electrodes in BESs. Such microorganisms are referred to as electroactive microorganisms. They possess the ability to attach and mediate electron transfer to the solid-state electron acceptors and from donors such as electrodes to sustain their respiratory or metabolic activities. BESs can be operated in two different modes, microbial fuel cell (MFC) and microbial electrolysis cell (MEC) depending upon the target process. Microbial substrate oxidation reaction at the anode is typically used for the treatment of organic matter containing wastewaters. Whereas substrate reduction reactions at the cathode are specifically used for the production of reduced products such as hydrogen (Rozendal et al., 2006) and biochemicals (Patil et al., 2015), or the remediation of recalcitrant contaminants and reduction of oxidized forms of heavy metals present in different kinds of wastes, process streams, wastewaters, and leachate solutions (Modin et al., 2012; He and Angenent, 2006). It can also be used for the reduction of other electron acceptors such as perchlorate, nitrobenzene, azo dyes, and nitrate (Ter Heijne et al., 2006; Strycharz-Glaven et al., 2013). In BESs, the electrons produced by the microbial oxidation of organic matter at the anode can be used to drive or facilitate the reduction of heavy metals at the cathode (Fig. 3.1). There are two possibilities for the

Removal of heavy metals using bioelectrochemical systems

57

Figure 3.1 Schematic of a two-chambered bioelectrochemical system (BES) showing the principle of heavy metals removal.

reduction of heavy metals at the cathode in BESs. First, if the reduction potential of the heavy metal at the cathode is higher (or more positive) than the oxidation potential of the electron donor at the anode, the reaction will proceed spontaneously and generate a net positive cell voltage in MFC mode operation. The metals that can be reduced at the cathode of MFCs include Cu21/Cu0, Se41/Se0, Cr61/Cr31, and V51/V41. The second possibility is through the use of MECs when the reduction potential of the metal is lower (or more negative) than the oxidation potential of electron donor; an external energy needs to be applied to facilitate the metal reduction reaction at the cathode. A few representative heavy metal reduction reactions that can occur at the cathode of BESs (both MFC and MEC) are presented below. 0 Ag1 ðaqÞ 1 e2 ! Ag0 ðsÞ; Ered 5 0:799 V versus SHE ðpH 7:0Þ (Choi and Cui, 2012) 22 1 2 0 Cr2 O7 ðaqÞ 1 6e 1 14H ! 2Cr31 ðaqÞ 1 7H2 O; Ered 5 1:33 V versus SHEðpH 2:0Þ (Li et al., 2008) 0 Co21 ðaqÞ 1 2e2 ! Co0 ðsÞ; Ered 5 2 0:232 V versus SHEðpH 2:0Þ (Huang et al., 2014a,b) 21 0 2 0 Cu ðaqÞ 1 2e ! Cu ðsÞ; Ered 5 0:34 V versus SHEðpH 7:0Þ (Modin et al., 2012)

where  SHE denotes standard hydrogen electrode, aq denotes aqueous, and ‘s’ denotes solid.

3.2.2 Reduction of heavy metals at the cathode of bioelectrochemical systems At the cathode of BESs, highly oxidized and unstable states of heavy metals get reduced by acquiring the electrons coming from the anode via external circuit. The majority of heavy metals can be reduced electrochemically at the abiotic cathode

58

Integrated Microbial Fuel Cells for Wastewater Treatment

(Velvizhi and Mohan, 2011; Mathuriya and Sharma, 2009). Examples include copper (Cu21/Cu20 at 0.340 V vs SHE) from the fly ash leachate by deposition (Tao et al., 2011; Ter Heijne et al., 2010) and Cr61/Cr31 (11.33 V vs SHE) reduction from metal-laden wastewater (Huang et al., 2010; Tandukar et al., 2009). The reduction of heavy metal is governed by several parameters such as the developed electrode potential at the cathode in case of MFC, number of electrons transferred per unit time, applied electric potential at the cathode in case of MEC, and the availability of other electron acceptors at the cathode (Modin et al., 2012). The use of microorganisms has also been reported for the heavy metal reduction at the cathode of BESs (He and Angenent, 2006; Ter Heijne et al., 2006). A brief description of microbial catalyzed heavy metal reduction is illustrated in Table 3.3. Mostly selected mixed microbial communities or pure cultures that are resistant to the metal toxicity and capable of utilizing a particular concentration of heavy metals for achieving their respiratory processes are used. For example, reduction of toxic Cr61 (by Shewanella decolorationis S12, Actinobacteria, Actinobacter, Geobacter, Shewanella oneidensis MR-1, Klebsiella pneumoniae L17, and mixed culture) and U61 (by Anaeromyxobacter dehalogenans, Clostridium sphenoides, Desulfotomaculum reducens, Desulfovibrio desulfuricans, Desulfovibrio sp., Geobacter metallireducens, Pseudomonas putida, Pseudomonas sp., Shewanella algae, S. oneidensis MR-1, Shewanella putrefaciens) to less toxic and less soluble products that can be easily recovered has been reported at the biotic cathodes of MFCs (Tandukar et al., 2009; Nancharaiah et al., 2010, 2016). The use of microorganisms for metal reduction reactions at cathode offers several advantages such as lowering the required electrochemical reduction potential, and recovery of economically value-added metallic products. The reduced heavy metals are either deposited at the cathode or settled in the reactor or form another compound by binding with any other chemical (Modestra et al., 2017). Thus in addition to the removal, BES also offers the possibility to recover some of the heavy metals in elemental or other forms. The bioelectrochemical removal of heavy metals depends on several factors. These include, for example, redox potentials of both half-cell reactions, initial metal ion concentration, the conductivity of the solution, the presence of cocontaminants or other competing electron acceptors, pH, and electrode materials. It also depends on the capability of metal ions to accept the electrons to get reduced (Rodrigo et al., 2010).

3.3

Electrode materials used for heavy metal removal in bioelectrochemical systems

The performance of BES is determined considerably by the type of electrode materials used. Though the electrodes are mainly utilized for redox processes in BESs, they also interact with the biocatalysts or microorganisms. In general, the electrode material should be conductive, stable, biocompatible, cost-effective, and nonhazardous to the environment. Generally, three different types of materials have been used

Table 3.3 An overview of various studies on the removal or recovery of heavy metals using bioelectrochemical systems (BESs). S. no.

Target metal

BES configuration, electrode materials

Electron donor conc.

Source of inoculum

Microbial culture

Maximum power/ current output for MFCs or applied voltage/potential in MECs

Reference

1.

Cr (VI)

Mixed anaerobic culture

150 mW/m2

Wang et al. (2008)

Cr (VI)

Sodium acetate, 2.64 g/L Acetate, 1.64 g/L

Anaerobic sludge

2.

Domestic wastewater

Mixed culture

1.6 mW/m2

Li et al. (2008)

3.

Cr (VI)

Anaerobic digester

Mixed culture

55.5 mW/m2

4.

Cr (VI)

Indigenous bacteria Actinobacteria

2.4 W/m3 or 6.9 A/m3

Cr (VI)

Shewanella oneidensis MR-1

32.5 mA/m2

Xafenias et al. (2013)

6.

NaVO3

MFC

Glucose

Contaminated soil from electroplating plant Artificial wastewater supplemented with Potassium chromate Metallurgical wastewater

Tandukar et al. (2009) Huang et al. (2010)

5.

MFC, graphite plates for anode and cathode MFC, graphite plate anode, rutile coated graphite plate cathode MFC, graphite plates as anode and cathode MFC, graphite plate anode, and granular graphite cathode Biocathodic MFC, graphite felt as electrodes

Rhodoferax ferrireducens

0.6 mA

Li et al. (2009)

Sodium acetate, 0.2 g/L Sodium acetate, 1.6 g/L Lactate and Acetate, 30 mM

(Continued)

Table 3.3 (Continued) S. no.

Target metal

BES configuration, electrode materials

Electron donor conc.

Source of inoculum

Microbial culture

Maximum power/ current output for MFCs or applied voltage/potential in MECs

Reference

7.

Au31

Trypticase soya agar

Anaerobic sludge

Shewanella putrefaciens

6.58 W/m2

8.

Se

Gold electrodeposition on G-10 graphite electrodes MFC

Glucose

Mine-waste calcines

S. oneidensis MR-1

13 mW/m2

9.

Se

Single-chambered MFC

Domestic wastewater

Mixed culture

1.500 mW/m2

10.

U (VI), (IV)

MFC, graphite rod electrodes

Sodium acetate and glucose Sodium acetate

DominguezBenetton et al. (2018) Nancharaiah et al. (2015) Catal et al. (2009)

Geobacter sulfurreducens PCA



Gregory and Lovley (2005)

11.

Co

MFC, graphite felt for anode

Contaminated soil from uranium ore processing facility Primary sludge from wastewater treatment plant

Abiotic cathode

Varied according to various factors in a range of 58 6 933 6 14 mW/ m3

Huang et al. (2013)

Sodium acetate, 0.38 gCOD/L

12.

Co

13.

Ni

14.

Ag

15.

Ag

16.

Ag

17.

Hg

18.

Cu

MEC, graphite brush anode, and graphite felt cathode MEC, carbon felt anode and stainless steel cathode MFC, carbon brush anode and carbon felt cathode MFC, graphite rod anode and graphite felt cathode MFC, carbon cloth anode and graphite felt cathode MFC, graphite felt anode and carbon paper cathode

Sodium acetate, 1 g/L

Primary MFC

Mixed culture

0.30.5 V

Jiang et al. (2014)

Sodium acetate, 1 g/L



Mixed culture

0.5 - 1.1 V

Qin et al. (2012)

Sodium acetate, 1 g/L Sodium acetate, 1.28 g/L

Sludge

Mixed culture

4.25 mW/m2

Laboratory-scale BES

Mixed culture

0.109 mW/m2

Choi and Cui (2012) Nancharaiah et al. (2015)

Sodium acetate, 1.6 g/L

Mixed culture

0.3 mW/m2

Yun-Hai et al. (2013)

Abiotic cathode

0.4333 mW/m2

Wang et al. (2011)

MFC, graphite plates electrodes

Glucose, 5 g/L

Anaerobic sludge from fruit waste treatment plant Anaerobic inoculum from wastewater treatment plant Anolyte sludge and catholyte CuSO4

Mixed culture

339 mW/m2

Tao et al. (2011)

Sodium acetate, 0.82 g/L

(Continued)

Table 3.3 (Continued) S. no.

Target metal

BES configuration, electrode materials

Electron donor conc.

Source of inoculum

Microbial culture

Maximum power/ current output for MFCs or applied voltage/potential in MECs

Reference

19.

Cu

Sodium acetate, 1.64 g/L

Running MFC anolyte

Mixed culture

0.43 mW/m2

Ter Heijne et al. (2010)

20.

Cd, Zn, Pb, Cu

Sodium acetate, 1.64 g/L

0 V—Cu 0.3 V—Pb 0.5 V—Cd 1.7 V—Zn 4.25 mW/m2

Modin et al. (2012)

Cu21

Mixture of aerobic and anaerobic sludge Excess sludge as anolyte and CuSO4 as catholyte

Mixed culture

21.

MFC, graphite plate anode and graphite foil pressed Ti-plate cathode MEC, carbon felt anode and titanium wire cathode Membrane-free baffled MFC, graphite plates as electrodes

MEC, Microbial electrolysis cell; MFC, microbial fuel cell.

Glucose, 5 g/L

Mixed culture

Guo et al. (2010)

Removal of heavy metals using bioelectrochemical systems

63

as electrodes in bioelectrochemical removal of heavy metals: carbon based, metal based, and composite materials (carbon 1 metals). Carbon-based electrode materials possess good biocompatibility and resistance to corrosion, which make them most suitable. However, some of the intrinsic characteristics such as low conductivity and low mechanical strength limit their use for large scale applications (Gupta and Joia, 2016; Modestra et al., 2017; Nancharaiah et al., 2015). Metallic electrodes, such as gold, platinum, titanium, stainless steel, copper, and aluminum, have good conductivity and mechanical strength but also bear some limitations such as the high cost and poor biocompatibility. For instance, copper is toxic to microbes, and stainless steel is cheap but has less biocompatibility. Electrode materials can be used in different configurations that may include both planar and three-dimension (3D) structures. The planar, veil, and mesh forms possess uniform surface properties that make the biofilm visualization and analysis easy. Though these are more suitable for fundamental research, recently, 3D electrodes are used to improve the microbial remediation processes as they provide comparatively more surface area for biofilm development and electrocatalysis. Along with the electrode type and configuration, the properties of electrodes such as surface topography and surface chemistry also influence the target remediation process (Guo et al., 2015). The selection of cathode materials for heavy metal reduction reactions should be done carefully based on its high specificity, electroactive surface, stability, high conductivity, fast electron transfer, high mechanical strength, good processability, low environmental impact, scalability, and the cost. For instance, carbon-based electrodes may support the adsorption of heavy metals forming a molecular or atomic film (the adsorbate) on it due to its high porosity and graphite-like microcrystalline lattice structure (Lakherwal, 2014). None among the carbon-based or metal electrodes can fulfill all of these criteria completely. So nowadays the combination of the best materials with specific surface modification is considered to generate composite electrode materials. The different types of cathode materials along with their advantages and disadvantages are presented in Table 3.4.

3.4

Conventional technologies versus bioelectrochemical systems-based technology for the removal of heavy metals

The development of novel, innovative, and site-specific techniques is imperative for the feasible and effective remediation of various heavy metalcontaminated sites. The conventional methodology to remediate heavy metals is dependent on physical, chemical, and biological approaches (Table 3.2). There is also a way to use the already existing technologies in concoction with each other to increase the remediation rate and to invent a novel and advanced methodology for the heavy metal remediation. Though they can remove the heavy metal contamination up to the desired

Table 3.4 Different types of electrode materials used in bioelectrochemical systems for the removal of heavy metals. S. no.

Type of electrode

Applications

Heavy metal(s) under test

Remarks

References

1.

Graphite plate

Used as both anode and cathode

Chromium, copper, and silver

High specific surface area, biocompatible

2.

Graphite felt

Used as both anode and cathode

Cobalt, copper, silver, and chromium

Possess high surface area but require tedious treatment

3.

Granular graphite

Cathode

Chromium

4.

Carbon cloth

Used as both anode and cathode

5.

Carbon brush

6.

Carbon fiber

7.

Carbon paper

Used as both anode and cathode Used as both anode and cathode Cathode

Copper, gold, silver, cadmium, and selenium Gold, silver, and cadmium Vanadium

Has sufficient surface area, but faces the problem of detached particles Acheive high coulombic efficiency and high current densities without separation

Wang et al. (2008) Tandukar et al. (2009) Ter Heijne et al. (2010) Wang and Ren (2014) Yun-Hai et al. (2013) Wang et al. (2010) Nancharaiah et al. (2015) Huang et al. (2014a, b) Huang et al. (2010) Nancharaiah et al. (2015) Nancharaiah et al. (2015) Choi and Cui (2012)

Mercury

8.

Carbon rod

Cathode

Cobalt

9.

Titanium plate with graphite foil Stainless steel

Cathode

Copper

Cathode

Nickel

10.

MEC, Microbial electrolysis cell; MFC, microbial fuel cell.

Improved power densities with mixed consortia of MFCs High surface area, high current densities for electroactive microbial biofilm Improved microbial electron transferbased processes Improved processes but have a lesser exposed surface area Good electrode material, provide good conductivity and sufficient surface area Possess good conductivity but faces the problem of oxidation on applying potential

Choi and Hu (2013) Nancharaiah et al. (2015) Zhang et al. (2009) Nancharaiah et al. (2015) (Wang et al. 2011) Nancharaiah et al. (2015) Jiang et al. (2014) Nancharaiah et al. (2015) Ter Heijne et al. (2010) Qin et al. (2012)

Removal of heavy metals using bioelectrochemical systems

65

level, they suffer due to various disadvantages and limitations. The physical techniques that can be applied at highly contaminated sites include soil replacement, soil isolation, vitrification, and electrokinetic methodology. These techniques need a lot of labor, which can produce hazardous waste, and are also costly to apply. Before the use of biological agents, it was believed that the use of chemicals for the removal and recovery of heavy metals is the best fitted and suitable approach because of their rapid activity and effectiveness at a wide range and concentration of heavy metals. However, they are also not so efficient for complete remediation because their activity depends on types of chemicals, soil, and metals. Some of the advanced conventional treatment processes such as membrane filtration and chemical precipitation may achieve maximum remediation rates (Kurniawan et al., 2006), but face limitations due to high cost and production of foul smell during the process. Nanofiltration is another effective technique to filter out or remove micropollutants, including heavy metals. However, it also exhibits the same limitations of fouling and pore chocking besides the high costs. Another promising conventional technique is the use of ionexchange resins where the charged metal ions are exchanged with the countercharged ions at the resins. Methods such as coagulation and flocculation can also be used but are not attractive due to the production of excess sludge (Rengaraj et al., 2001; Andrus, 2000; Modestra et al., 2017). The shift of heavy metal contaminants from one source to another is also a major disadvantage of conventional methods. Some other limitations include their lesser selectivity, sensitivity, and specificity. Most of them also need pretreatment processes. Heavy metal removal using biological approaches is considered economically suitable and eco-friendly. Besides microbial processes, some of the useful plant-based methods include phytoextraction, phytostabilization, and phytovolatilization. Besides being time-consuming, they suffer due to limitations of restricted use and low efficiency at moderate conditions. Biological remediation techniques have advantages of removal even at mild conditions or in low nutrient supply along with the production of differently desired by-products using the nutrients and growth-supportive elements from wastes. However, they also bear the limitations such as the need for additional electron donors and acceptors to complete the respiration process. Although these conventional methods are used widely, they are not considered successful for heavy metal removal if the cost-effectiveness, limitations, and harmful or xenobiotic products generation are considered (Wuana and Okieimen, 2014). In this context, the bioelectrochemical approach is poised to play an important role in the future development of heavy metal removal and recovery technologies. It is possible to use BESs for the continuous checking and monitoring of the microbial activities and metabolic functions along with a constant supply of electronaccepting or -donating conditions. For example, in the case of reductive removal of oxidized metals such as mercury, lead, and copper, the cathode can be used as the nonlimiting source of electron donor. It can also address the problems of external chemical use, contamination, incomplete waste degradation, and lesser by-product formation (Aulenta et al., 2008). Both biotic and abiotic cathods are efficient for the removal and recovery of heavy metals. Process efficiency of the abiotic cathode is more compared to the biotic cathodic approach as it does not interrupt in the

66

Integrated Microbial Fuel Cells for Wastewater Treatment

Figure 3.2 Advantages of using BES approach for the removal of heavy metals. BES, Bioelectrochemical system.

recovery of metals from treated effluents, but the required electrode potential may be higher in this case. The development of biofilm reduces the requirement of potential at the biotic cathode. Also, the reduction reactions by the biotic cathode are eco-friendly and have the ability of biocatalyst self-regeneration. Other special characteristics such as faster reduction rate, low price, and mild operational conditions make it more advantageous over the abiotic cathode. Hence, it can be preferred and advantageous for heavy metal removal. Some of the major advantages of BES technology are listed in Fig. 3.2 (Kaushik, 2015). The BES technology has also some limitations mainly due to lack of inadequate fundamental understanding of electron transfer mechanisms, lack of efficient and cheaper electrode and separator materials, and requirement of process monitoring and control. It is a tedious process to develop microbial biofilms at the electrodes, as it requires a proper supply of medium, specific environment, and conditions. Continuous monitoring is also required to maintain the active microorganisms in BESs. It is a slow process in comparison to the electrochemical approach. In the case of MEC-based processes, it requires some external power supply to facilitate nonspontaneous metal reduction reactions at the cathode. A lot of fundamental and engineering research is thus required in advancing the BES technology as the promising approach to remove different kinds of pollutants along with heavy metals from the contaminated environments.

3.5

Conclusion

This chapter provides a brief overview of heavy metal contamination, associated problems, and their removal using different technologies. Among various

Removal of heavy metals using bioelectrochemical systems

67

technologies, BES is an emerging and powerful technology for removing different types of pollutants including heavy metals from contaminated environments. It is still in the development stage, and thus more research is needed to be done to understand the microbemetal interactions, electron transfer mechanisms, and electrode materials to enhance the efficiency of bioelectrochemical processes. The future research should also be focused on the recovery of usable forms of heavy metals from wastewaters and other waste streams.

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Gregory, K.B., Lovley, D.R., 2005. Remediation and recovery of uranium from contaminated subsurface environments with electrodes. Environ. Sci. Technol. 39, 89438947. Guo, H., et al., 2010. Bioremediation of heavy metals by growing hyperaccumulator endophytic bacterium Bacillus sp. L14. Bioresour. Technol. 101 (22), 85998605. Guo, K., et al., 2015. Engineering electrodes for microbial electrocatalysis. Curr. Trends Biotechnol. 33, 149156. Gupta, A., Joia, J., 2016. Microbes as potential tool for remediation of heavy metals: a review. J. Microb. Biochem. Technol. 8, 364372. He, Z., Angenent, L.T., 2006. Application of bacterial biocathodes in microbial fuel cells. Electroanalysis 18, 20092015. Hernberg, S., 2000. Lead poisoning in a historical perspective. Am. J. Ind. Med. 38, 244254. Huang, L., Chen, J., Quan, X., Yang, F., 2010. Enhancement of hexavalent chromium reduction and electricity production from a biocathode microbial fuel cell. Bioprocess. Biosyst. Eng. 33, 937945. Huang, L., et al., 2013. Synergetic interactions improve cobalt leaching from lithium cobalt oxide in microbial fuel cells. Bioresour. Technol. 128, 539546. Huang, L., et al., 2014a. Cobalt recovery with simultaneous methane and acetate production in biocathode microbial electrolysis cells. Chem. Eng. J. 253, 281290. Huang, L., Yao, B., Wu, D., Quan, X., 2014b. Complete cobalt recovery from lithium cobalt oxide in self-driven microbial fuel cell  microbial electrolysis cell systems. J. Power Sources 259, 5464. Hutton, M., 1987. Human health concerns of lead, mercury, cadmium, and arsenic. Lead, Mercury, Cadmium & Arsen. Environ. pp. 5368. Jacobs, I.A., et al., 2002. Poisoning as a result of barium styphnate explosion. Am. J. Ind. Med. 41, 285288. Jiang, L., Huang, L., Sun, Y., 2014. Recovery of flakey cobalt from aqueous Co(II) with simultaneous hydrogen production in microbial electrolysis cells. Int. J. Hydrogen Energy 39, 654663. Jobin, L., Namour, P., 2017. Bioremediation in water environment: controlled electrostimulation of organic matter self-purification in aquatic environments. Adv. Microbiol. 07, 813852. Kaushik, G. (Ed.), 2015. Applied Environmental Biotechnology: Present Scenario and Future Trends. Springer, India (Chapter 10). Khan, N., Camille, A., 2017. Open Archive Toulouse Archive Ouverte, vol. 182. pp. 247268. Kurniawan, T.A., et al., 2006. Physico-chemical treatment techniques for wastewater laden with heavy metals. Chem. Eng. J. 118 (12), 8398. Available from: https://doi.org/ 10.1016/j.cej.2006.01.015. Lakherwal, D., 2014. Adsorption of heavy metals: a review. Int. J. Environ. Res. Dev. 4 (1), 4148. Li, Z., Zhang, X., Lei, L., 2008. Electricity production during the treatment of real electroplating wastewater containing Cr6 1 using microbial fuel cell. Process Biochem. 43, 13521358. Li, Y., et al., 2009. Cr(VI) reduction at rutile-catalyzed cathode in microbial fuel cells. Electrochem. Commun. 11, 14961499. Maas, P.V., et al., 2005. Enzymatic versus nonenzymatic conversions during the reduction of EDTA-chelated Fe (III) in BioDeNOx reactors. Environ. Sci. Technol. 39, 26162623.

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Maher, W., Butler, E., 1988. Arsenic in the marine environment. Appl. Organomet. Chem. 2, 191214. Mathuriya, A.S., Sharma, V.N., 2009. Bioelectricity production from various wastewaters through microbial fuel cell technology. J. Biochem. Technol. 2, 133137. Minnikova, T.V., et al., 2016. Assessing the effect of heavy metals from the Novocherkassk power station emissions on the biological activity of soils in the adjacent areas. J. Geochem. Explor. 174. Modestra, J.A., et al., 2017. Bioelectrochemical systems for heavy metal removal and recovery. In: Rene, R.E., et al., (Eds.), Sustainable Heavy Metal Remediation. Springer International Publishing AG. Modin, O., Wang, X., Wu, X., Rauch, S., Fedje, K.K., 2012. Bioelectrochemical recovery of Cu, Pb, Cd, and Zn from dilute solutions. J. Hazard. Mater. 235236, 291297. Mohanty, M., Patra, H.K., 2012. Effect of chelate-assisted hexavalent chromium on physiological changes, biochemical alterations, and chromium bioavailability on crop plants  an in vitro phytoremediation approach. Biorem. J. 6 (3), 147155. Mukherjee, B., et al., 2004. Vanadium  an element of atypical biological significance. Toxicol. Lett. 150, 135143. Nancharaiah, Y.V., Lens, P.N.L., 2015. Selenium biomineralization for biotechnological applications. Trends Biotechnol. 33 (6), 18. Nancharaiah, Y.V., et al., 2010. Immobilization of Cr(VI) and Its reduction to Cr(III) phosphate by granular biofilms comprising a mixture of microbes. Appl. Environ. Microbiol. 76, 24332438. Nancharaiah, Y.V., Venkata Mohan, S., Lens, P.N.L., 2015. Metals removal and recovery in bioelectrochemical systems: a review. Bioresour. Technol. 195, 102114. Nancharaiah, Y.V., Mohan, S.V., Lens, P.N.L., 2016. Biological and bioelectrochemical recovery of critical and scarce metals. Trends Biotechnol. 34, 137155. Nanda, K., 1995. Phytoextraction: the use of plants to remove heavy metals from soils. Environ. Sci. Technol. 29 (5), 12321238. Patil, S.A., et al., 2015. Selective enrichment establishes a stable performing community for microbial electrosynthesis of acetate from CO2. Environ. Sci. Technol. 49 (14), 88338843. Pedersen, A.J., Ottosen, L.M., Villumsen, A., 2003. Electrodialytic removal of heavy metals from different fly ashes influence of heavy metal speciation in the ashes. J. Hazard. Mater. 100, 6578. Perfus-Barbeoch, L., Leonhardt, N., Vavasseur, A., Forestier, C., 2002. Heavy metal toxicity: cadmium permeates through calcium channels and disturbs the plant water status. Plant J. 32, 539548. Qin, B., et al., 2012. Nickel ion removal from wastewater using the microbial electrolysis cell. Bioresour. Technol. 121, 458461. Rengaraj, S., Yeon, K.H., Moon, S.H., 2001. Removal of chromium from water and wastewater by ion exchange resins. J. Hazard. Mater. 87, 273287. Richards, A., et al., 2007. C-terminal truncations in human 3’-5’ DNA exonuclease TREX1 cause autosomal dominant retinal vasculopathy with cerebral leukodystrophy. Nat. Genet. 39, 10681070. Rodrigo, M.A., Can˜izares, P., Lobato, J., 2010. Bioresource technology effect of the electron-acceptors on the performance of an MFC. Bioresour. Technol. 101, 70147018. Rozendal, R.A., et al., 2006. Principle and perspectives of hydrogen production through biocatalyzed electrolysis. Int. J. Hydrogen Energy 31, 16321640.

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Singh, A., Yakhmi, M.J.V., 2014. Microbial fuel cells to recover heavy metals. Environ. Chem. Lett. 12, 483494. Srivastava, N.K., Majumder, C.B., 2008. Novel biofiltration methods for the treatment of heavy metals from industrial wastewater. J. Hazard. Mater. 151, 18. Strycharz-Glaven, S.M., et al., 2013. Electrochemical investigation of a microbial solar cell reveals a nonphotosynthetic biocathode catalyst. Appl. Environ. Microbiol. 79, 39333942. Tandukar, M., Huber, S.J., Onodera, T., Pavlostathis, S.G., 2009. Biological chromium(VI) reduction in the cathode of a microbial fuel cell. Environ. Sci. Technol. 43, 81598165. Tao, H.C., et al., 2011. Removal of copper from aqueous solution by electrodeposition in cathode chamber of microbial fuel cell. J. Hazard. Mater. 189, 186192. Ter Heijne, A., et al., 2006. A bipolar membrane combined with ferric iron reduction as an efficient cathode system in microbial fuel cells. Environ. Sci. Technol. 40, 52005205. Ter Heijne, A., et al., 2010. Copper recovery combined with electricity production in a microbial fuel cell. Environ. Sci. Technol. 44, 43764381. Velvizhi, G., Mohan, S.V., 2011. Bioresource technology biocatalyst behavior under selfinduced electrogenic microenvironment in comparison with anaerobic treatment: evaluation with pharmaceutical wastewater for multi-pollutant removal. Bioresour. Technol. 102, 1078410793. Wang, G., Huang, Æ.L., Zhang, Æ.Y., 2008. Cathodic reduction of hexavalent chromium [Cr (VI)] coupled with electricity generation in microbial fuel cells. Biotechnol. Lett. 30, 19591966. Wang, H., Ren, Z.J., 2014. Bioelectrochemical metal recovery from wastewater: a review. Water Res. 66, 219232. Wang, Z., Lim, B., Lu, H., Fan, J., Choi, C., 2010. Cathodic reduction of Cu21 and electric power generation using a microbial fuel cell. Bull. Korean Chem. Soc. 31, 20252030. Wang, Z., Lim, B., Choi, C., 2011. Bioresource technology removal of Hg21 as an electron acceptor coupled with power generation using a microbial fuel cell. Bioresour. Technol. 102, 63046307. Watts, J., 2003. Concern over mercury pollution in India. Sci. Direct 362. Wu, G., et al., 2012. Short communication Hazardous heavy metal distribution in Dahuofang catchment, Fushun, Liaoning, an important industry city in china: a case study. Clean  Soil Air Water 40 (12), 13721375. Wuana, R.A., Okieimen, F.E., 2014. Heavy metals in contaminated soils: a review of sources, chemistry, risks, and best available strategies for remediation. Int. Sch. Res. Network ISRN Ecol. 2011, 150. Xafenias, N., Zhang, Y., Banks, C.J., 2013. Enhanced performance of hexavalent chromium reducing cathodes in the presence of Shewanella oneidensis MR-1 and lactate. Environ. Sci. Technol. 47, 45124520. Yun-Hai, W., et al., 2013. Electricity production from a bio-electrochemical cell for silver recovery in alkaline media. Appl. Energy 112, 13371341. Zhang, W., et al., 2009. Heavy metal contamination in surface sediments of Yangtze River intertidal zone: an assessment from different indexes. Environ. Pollut. 157, 15331543.

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Further reading Fredrickson, J.K., et al., 2008. Towards environmental systems biology of Shewanella. Nat. Rev. Microbiol. 6, 592603. Mahadevan, R., Palsson, B., Lovley, D.R., 2011. In situ to in silico and back: elucidating the physiology and ecology of Geobacter spp. using genome-scale modeling. Nat. Rev. Microbiol. 9, 3950. Mohan, S.V., Srikanth, S., 2011. Bioresource technology enhanced wastewater treatment efficiency through microbially catalyzed oxidation and reduction: synergistic effect of biocathode microenvironment. Bioresour. Technol. 102, 1021010220. Nealson, K.H., Scott, J., 2006. Ecophysiology of the genus Shewanella. In: Dworkin, M., et al., (Eds.), The Prokaryotes. Springer, New York. Patil, S.A., H¨agerh¨all, C., Gorton, L., 2012. Electron transfer mechanisms between microorganisms and electrodes in bioelectrochemical systems. Bioanal. Rev. 4, 159192. Singh, D.D., 2018. Next-generation sequencing technologies as emergent tools and their challenges in viral diagnostic. Biomed. Res. 29 (8), 16371644. Tao, H.C., et al., 2012. Recovery of silver from silver(I)-containing solutions in bioelectrochemical reactors. Bioresour. Technol. 111, 9297.

Textile wastewater treatment using microbial fuel cell and coupled technology: a green approach for detoxification and bioelectricity generation

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Supriya Gupta1, Yamini Mittal1, Prashansa Tamta2, Pratiksha Srivastava3 and Asheesh Kumar Yadav1 1 Academy of Scientific and Innovative Research (AcSIR), CSIR-Human Resource Development Centre, (CSIR-HRDC) Campus, Ghaziabad, India; Environment and Sustainability Department, CSIR-Institute of Minerals and Materials Technology, Bhubaneswar, India, 2University School of Environment Management, Guru Gobind Singh Indraprastha University, New Delhi, India, 3Australian Maritime College, College of Sciences and Engineering, University of Tasmania, Launceston, TAS, Australia

Chapter Outline 4.1 Microbial fuel cell and its application in the treatment 74 4.1.1 Mechanisms involved in dye breakdown 74 4.1.2 Dye removal and current generation in microbial fuel cell 76 4.1.3 Dye removal and total COD removal 77

4.2 Enhancement of microbial fuel cell performance 4.2.1 4.2.2 4.2.3 4.2.4

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Bioanode-based enhancement of dye treatment 78 Biocathode-based enhancement of dye treatment 79 Membrane-based enhancement of dye treatment 80 Effect of the shuttle on dye removal and electricity generation 81

4.3 Microbial diversity involved in the breakdown of dye in microbial fuel cell 81 4.4 Toxicity of treated dye wastewater 82 4.5 Microbial fuel cell
coupled techniques for textile wastewater treatment 83 4.5.1 4.5.2 4.5.3 4.5.4

Microbial fuel cell
integrated constructed wetlands 83 Microbial fuel cell couple aerobic biocontact oxidation reactor system 84 Bioelectro-Fenton technology-microbial fuel cell 85 Electrolysis cell combined with a microbial fuel cell (MFC-MEC) 85

4.6 Research gap 86 Acknowledgments 87 References 87 Further reading 91

Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00004-7 © 2020 Elsevier Inc. All rights reserved.

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Microbial fuel cell and its application in the treatment

A microbial fuel cell (MFC) is one of the bioelectrochemical systems, exploiting microorganisms as a biocatalyst to oxidize organic and inorganic substrate in an anaerobic environment to generate electricity (Logan et al., 2006). It is a unique device that not only generates power but treats wastewater efficiently by catalyzing the oxidation and reduction reactions. Detail discussion is already given in the introductory chapters of this book. Lately, MFC has emerged out as a potential low-cost wastewater treatment technology that shows its potential in the treatment of textile wastewater, especially containing azo dye. Azo dyes are the aromatic compounds with one or more
N 5 N
groups, known as a chromophore. A chromophore group is a delocalized electron system with conjugated double bonds, which is responsible for light absorption in dye molecules (Ali, 2010). In the case of azo dyes, chromophore (
N 5 N
group) attributes to its recalcitrant nature and restricts its bio- and/or photodegradation. The presence of sulfonate groups contributes equally to it. These groups impart electron-withdrawal character to dye molecule that generates electron deficiency and makes azo compounds less susceptible to oxidative catabolism. Generally, azo dyes are treated by physical or chemical methods, but these are tagged with certain limitations. The physical treatment involves coagulation
flocculation, electrocoagulation, ion exchange, membrane filtration, etc. These techniques are associated with excess sludge generation, handling and disposal problems, and high operating costs. Several chemical treatments such as application of sodium hypochlorite, redox-active metals, photochemical treatment, electrochemical destruction, and ozonation also deal with sludge production, generation of toxins and carcinogens, etc. (Amaral et al., 2004; Bafana et al., 2011). MFC presents an appropriate solution to the aforementioned problems. It reduces sludge generation, CO2 emission, and operational cost as compared to the conventional aerobic and anaerobic process. Moreover, MFC enhances the decolorization of azo dyes and at the same time, it recovers electricity from a readily degradable organic matter present in the wastewater stream. In recent years, MFC has been extensively studied for textile wastewater decolorization, treatment, and detoxification (Mu et al., 2009; Cao et al., 2017; Sun et al., 2016).

4.1.1 Mechanisms involved in dye breakdown Azo dyes, most commonly used synthetic dyes, are heterocyclic aromatic compounds linked with color display and polar groups. Their degradation mechanism involves the reduction of azo bonds (N 5 N) in the presence of some electron donors to form simpler aromatic amines that can be further oxidized for the complete mineralization of dye. Similar reductive condition prevails in MFC where the anodic and cathodic chambers, separated by a membrane, are marked by anaerobic (oxidative) and aerobic (reductive) reactions, respectively. At the anode, electrons are generated during the oxidation of the substrate (carbon source) by

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electrochemically active microorganisms. Subsequently, the microorganisms transfer these electrons to the external electron acceptors via extracellular electron transfer. When the artificial electron acceptors/donors such as electrodes in two chambers are connected externally through a wire, the electrons start to flow from anode to cathode through the circuit, thus producing current. Protons, generated during oxidation, transfer through the proton exchange membrane (PEM) to the cathode, where these combine with electron and oxygen to form water. This basic principle of MFC benefits dye removal by intensive reduction and oxidation reactions occurring altogether in one system. The cleaving of azo bonds by electrons liberated from cosubstrate oxidation in an anaerobic chamber (anode) reduces the long, complex dye structure to simpler aromatic amines efficiently. Furthermore, the combination of electrons (transiting through external circuit) with azo dye in cathodic chamber intensifies the azo bond (N 5 N) reduction with the formation of colorless and biodegradable aromatic amines and sulfanilic acids (Frijters et al., 2006; Van der Zee et al., 2001). The aerobic environment of the MFC cathode assists the oxidation of aromatic amines for complete mineralization and detoxification. In general, at anode azo bonds break biologically under anaerobic conditions and at intermediate cathode products degrade and mineralize (Li et al., 2010) (Fig. 4.1). Here, decolorization of azo dyes is possible because of biodegradation rather than biosorption of bacterial cells (Sun et al., 2009a). With the function of operational parameters and other favorable factors, diverse and dynamic microbial community outcompetes in an MFC system. Such microbial diversity with varying biochemistry degrades azo dyes either by intracellular or

Figure 4.1 Structure of three common azo dyes and their reduced products formed in MFC. MFC, Microbial fuel cell.

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extracellular metabolic reactions. The intracellular reduction of azo dye requires the presence of an azoreductase enzyme with a specific transport system for the uptake of dye for its reduction inside the cell. In extracellular reduction the dye might act as an electron acceptor for electrons supplied via the carriers of electron transport chain or by the reduced compounds generated from anaerobic biomass (Ilamathi and Jayapriya, 2018). Chemical structures of dyes also affect the rate of their decolorization. The treatment efficiency and electric production decrease with the presence of mono-, di-, and poly azo bonds in the same order. Hsueh et al. (2009) reported that azo bond with a less electron density becomes more electrophilic for reductive biodecolorization. Also, the electron-withdrawing group is easily degradable than the electronreleasing group, and the position of the electron-withdrawing group to azo bond (N 5 N) significantly affects the rate of decolorization. The rate of decolorization is higher in the para-position than ortho- and meta-positions. The lowest rate of decolorization can be observed in meta-position because para- and ortho-position can more efficiently withdraw electron from azo bond through resonance. Hence, the azo bond becomes more electrophilic to be reduced for decolorization (Solanki et al., 2013; Oon et al., 2017).

4.1.2 Dye removal and current generation in microbial fuel cell The simultaneous dye decolorization, detoxification, and current generation have been studied rigorously in MFC systems. The relations between dye concentration in influent, removal in the effluent, and electricity production have been focused on enhanced performance efficiency. The decolorization rate and electricity generation are found dependent on many factors such as dye type and structure, concentration of the dye used, operating conditions such as pH and hydraulic retention time (HRT), type of wastewater, electrode material, and external resistance as a whole (Solanki et al., 2013; Varanasi and Das, 2018; Hou et al., 2011, 2012). The presence of dye in the anode of MFC presents competition to anode electrode for electron capture, and the power output tends to decline gradually. Thus with an increase in the dye concentration, the columbic efficiency of the system decreases (Sun et al., 2009a, 2013a). The dye concentration is also known to slightly decrease the catalytic activity of anodic microbes (including exoelectrogens), which further affects the electron generation at anode adversely. But the anode (electron acceptor) further facilitates and promotes the growth of biofilm to stabilize the voltage gradually (Sun et al., 2009a). Kalathil et al. (2011) studied the dye removal from real dye wastewater in granular activated carbon (GAC)-based MFC and reported 73.0% and 77.0% color removal at anode and cathode, respectively, with a power density of 1.7 W/m3 at 2 days HRT. The GAC particles as electrodes provided high surface area for enhanced biofilm formation, a key factor for high activity of MFC that was responsible for enhanced performance. Similarly, Li et al. (2010) demonstrated Congo red decolorization using graphite granules as a cathode. They reported the maximum power density of 364.50 mW/m2 with a small external resistance (50 Ω) and a total

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decolorization efficiency of 96.0% with an HRT of 45 h. The granules in both studies provided high surface area for cathodic reduction reaction and enhanced electricity generation. Electricity production could be significantly affected by the redox potential of the substrate. Cao et al. (2010) examined the effect of cosubstrate on Congo red dye decolorization and electricity generation in PEM air-cathode single-chamber MFC. They studied glucose, acetate, and ethanol separately as cosubstrates and found more than 98.0% dye decolorization in 36 h with maximum power density of 103 mW/m2 for glucose compared with 85.9 mW/m2 for acetate and 63.2 mW/m2 for ethanol. Thus, they concluded that electricity production could be significantly affected by the redox potential of the substrate. The fastest decolorization was reported in glucose, followed by ethanol than acetate. Niu et al. (2012) reported an enhanced power output and color-removal efficiency of Orange G (OG) dye while treated in MFC, modified the cathode with Fe (II)-EDTA (ethylene-diamine-tetra-acetic acid) catalyzed persulfate as the cathode solution and under optimal conditions obtained the maximum power density of 91.1 mW/m2 with OG removal rate 97.4% in 12 h. They concluded that EDTA could improve the stability of voltage output.

4.1.3 Dye removal and total COD removal As aforesaid, the chemical oxygen demand (COD) present in wastewater may act as the electron donor for the reduction of azo dyes. Thus, the dye removal is mainly dependent on its type and concentration. Glucose is known to be the best cosubstrate for azo dye decolorization (Cao et al., 2010; Li et al., 2010). Li et al. (2010) stated that increased concentration of glucose from 100 to 4000 mg/L increased the Congo red decolorization from 42.7% to 77.0% in the anodic chamber, whereas the overall decolorization increased from 69.3% to 92.7%. The dye concentration in wastewater affects the overall COD removal. With increasing concentrations of dyes, overall COD removal decreases. This can be due to the adverse effect of dye on the catalytic activity of anodic biofilm that oxidized the cosubstrate (COD). Sun et al. (2009a) reported the overall decrement in COD removal from 100% to 60% when dye concentration increased from 200 to 1500 mg/L, and organic load increased from 525 to 800 mg/L. However, at the same time, they concluded that dye decolorization was not strongly inhibited even at higher concentrations of 1500 mg/L, and removal efficiency was still observed up to 77.0% in 48 h HRT. This confirmed that the decolorization activity of biocatalysts affected slightly, and the anodic microbial consortium has excellent toxic tolerance for dye. Furthermore, dye reduces to form simpler amines and sulfanilic compounds. These residues and their degraded products contribute to COD as well that indeed shows decreased overall COD removal (Sun et al., 2009a). In other similar studies with dye wastewater, Kalathil et al. (2012) achieved 71.0% COD removal and 75.0% decolorization in 48 h of HRT; and maximum power density of 8 W/m3 with an open-circuit voltage (OCV) 0.39 V using a modified single-chamber MFC system with GAC biocathode.

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Enhancement of microbial fuel cell performance

The design of MFCs is continuously improving to enhance its performance for bringing it up to field scale or practical scale. Some limitations, such as overpotential, membrane and concentration resistance, and other internal resistances that hamper the performance of MFC are continuously being worked with (Gil et al., 2003). For enhancement of MFC performance with respect to dye removal, some modifications have been made in cathode and anode electrodes, and cationexchange membrane. The economic feasibility of systems has also been considered at the same time.

4.2.1 Bioanode-based enhancement of dye treatment More recently, the bioelectrode systems such as bioanodes and biocathodes have been found useful for the removal of dyes and many more recalcitrant contaminants present in the effluent from dye manufacturing and consuming industries. Sun et al. (2009a) found accelerated decolorization of azo dye (active brilliant red X-3B) in an air-cathode single-chambered MFC and generated electricity in the presence of readily biodegradable organic matters. But at the same time, the presence of azo dye adversely affected the thickness of biofilm formed. The scanning electron microscope images of anode electrodes revealed that thick biofilm developed in the absence of Congo red dye, in contrast to the sparse biofilm, developed in its presence. Sun et al. (2011) found a delay in the maximum stable voltage output due to slowly developed anode potential during Congo red decolorization. The reason can be totally understood by the mechanism of dye degradation. The electrons generated during the oxidation of substrate are preferred for the utilization of dye degradation or decolorization rather than electricity generation. The cyclic voltammetry results suggested that Congo red is a more favorable electron acceptor than the anode itself. Meanwhile, it has also been reported that azo dye and its decolorized products can act as a shuttle for the transfer of electrons from the bacterial surface to the anode at low concentration but resulted in accelerated consumption of electrons at high concentration. Thus, the application of bioanode prevents the retarded biofilm growth phase and contributes to the enhancement of azo dye degradation and electricity generation. De Dios et al. (2013) exploited the synergy between the Trametes versicolor (fungi) and Shewanella oneidensis MR-1 (bacteria) by acclimating the graphite with fungal
bacterial culture (bioanode) for decolorization of lissamine green B and crystal violet. They concluded that such symbiosis reduces the electron transport barriers from the bacteria to the anode since the hyphae framework of fungi is utilized by bacteria for growth, which is responsible for dye degradation enhancement. They reported a maximum volumetric power density of 0.78 W/m3, which is highest achieved yet. Guo et al. (2014) worked on the surface modification of anode to understand biofilm formation while treating methyl orange (MO) dye wastewater. The anode

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was fabricated with carbon paper (CP) and graphene (GR) via layer-by-layer assembly technique to promote electricity generation and MO dye removal and found a stable maximum power density of 368 mW/m2 with 1000 Ω external resistance. Through scanning electron microscopy, they revealed that surface roughness of CP/GR had been increasing, and due to this, it became favorable for more attachment of bacteria on the anodic surface.

4.2.2 Biocathode-based enhancement of dye treatment The reduction of oxygen on the surface of carbon-based conductive material (cathode) is a kinetically slow process. It requires the catalytic action of either chemicals such as ferricyanide/hexaferrocyanate (Rabaey and Verstraete, 2005) or permanganate (You et al., 2006) or metals such as platinum or metal-based compounds such as lead dioxide, manganese dioxide molybdenum/vanadium, cobalt, and iron-based materials (Habermann and Pommer, 1991). Use of such catalysts has shown performance enhancement of MFC, but at the same time, chemicals used were toxic for anodic microorganisms, also aesthetically impermissible due to the high risk of environment pollution (Sun et al., 2011). Metal-based catalysts can entail inactivation as these are generally susceptible to the adverse environmental conditions that may occur in MFCs (Zhang et al., 2009). The replacement of chemicals at the cathode for oxygen reduction by incorporation of biocathode as a catalyst is useful in mitigating such issues. Sun et al. (2011) introduced a novel method for azo dye (ABRX3) removal accompanied by additional power output. In this method an aerobic biocathode is introduced, which eliminates the use of chemicals and metal-based catalysts. The anaerobic and aerobic biocathode can be classified on the basis of the final electron acceptor. In aerobic biocathode, the final electron acceptor is oxygen, and in the absence of oxygen, it can be nitrate, sulfate, iron, manganese, etc. (Stams et al., 2006). Simultaneous oxidation and reduction of redox couples Mn21 /Mn41(Rhoads et al., 2005) and Fe21/Fe31 (TerHeijne et al., 2007) lead to oxygen reduction which further is a reason for decolorization of azo dye-containing wastewater. The aerobic biocathode incorporated in MFC was found to remove the aromatic amine compounds produced during the reduction of azo dye regardless of the power output as was revealed from both open and closed circuit studies. An enzymatic biocathode using laccase was studied in dual-chamber MFC to enhance the removal of Reactive Blue 221 (RB221) dye without the application of external mediators (Bakhshian et al., 2011). It was found that laccase could catalyze the removal of RB221 and had a positive effect on MFC performance as well. The maximum power density also increased by 30.0% when enzymatic decolorization was performed in the cathode chamber. Kalathil et al. (2012) designed a GAC biocathode assembled with GAC-based single-chamber MFC (GACB-SCMFC) for decolorization of dye wastewater and electricity production. GACB-SCMFC treated effluent was much less toxic as compared to the initial dye-containing wastewater influent; it indicates that GAC biocathode can be a good alternative to platinum and other chemical catalysts. With

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the use of GAC biocathode, they even made a way to the most challenging limiting factor (i.e., electrode cost) for MFC with the advantage that pH was automatically adjusted to 8 in 48 h of HRT.

4.2.3 Membrane-based enhancement of dye treatment The membrane in any MFC is a crucial component. It is employed to separate the anodic chamber from the cathodic chamber physically and affects the overall treatment and power output performance of the reactor. Its application enhances the reactor performance by restricting oxygen diffusion into the anodic chamber which otherwise can lower down the electro catalytic activity of the anodic microorganisms (Liu and Logan, 2004) and preventing the direct contact of cathode catalyst with polluted environment of anode which can lead to the fouling and deterioration of the MFC performance (Tartakovsky and Guiot, 2006), whereas the concentration gradient develops across it which can increase internal resistance and decrease the coulombic efficiency of MFC. To deal with limitations of membranes, various types of membranes have been explored to find PEM as a most suitable candidate due to its high conductivity for cations/protons and lower internal resistance compared to another ultrafiltration and anion exchange membrane (Liu and Logan, 2004). Its large-scale usability is still restricted due to high cost, which can account for up to 40% of the total cost of MFC (Rahimnejad et al., 2014). Also, it is subjected to fouling in the long run, which adds to the maintenance and operation cost of MFC (Behera et al., 2010). So now, membranes are used as an optional component in MFC to normalize the whole system, that is, the cost as well as the performance. The first study on simultaneous azo dye decolorization and electricity generation using MFC was conducted by Sun et al. (2009a), where they employed microfiltration membrane (MFM) instead of PEM as an alternative for cost reduction. Although it was highly efficient in isolating the cathodic catalyst from anodic microorganisms, it posed a danger for decaying of the cathodic catalyst due to its high accessibility for soluble pollutants such as azo dye and degraded dye products (Sun et al., 2013a) that can reduce the power output. The diffusion of oxygen from cathode to anode is also not administered well by MFM that contributes to decreased coulomb recovery but enhanced oxidation of reduced intermediates and so dye removal. Membrane-less MFC then proved its potential as a better system than those employing MFM (Wang et al., 2013; Sun et al., 2009b; Zhang et al., 2015; Thung et al., 2015). It showed comparable performance without the cost of the expensive membrane. At a current density of 0.552 A/m2, the MFC with an MFM achieved maximum power densities, 214 mW/m2 and 878 mW/m3, which were comparable to those of the membrane-less MFC, 208 mW/m2 and 819 mW/m3 (Sun et al., 2009b). Further studies by researchers continued on membrane-less MFC for dye removal but with modifications in the designs and operating parameters. Kalathil et al. (2011) used GAC as electrodes in membrane-less MFC without the application of platinum as cathode catalyst and removed substantial amount of dye and COD from wastewater along with the simultaneous maximum power density of

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1.7 W/m3 at an open-circuit voltage of 0.45 V. The cathode effluent was found to be almost nontoxic. Cui et al. (2012) combined the membrane-free upflow biocatalyzed electrolysis reactor with aerobic biocontact oxidation reactor (ABOR) and supplied the external power of 0.5 V to attain enhanced decolorization efficiency up to 94.8% 6 1.5% for Alizarin yellow in just 2 h HRT with the AYR loading rate of 780 g/m3 TV/day. The coupled oxidation reactor completely mineralized the dye reduced products.

4.2.4 Effect of the shuttle on dye removal and electricity generation As illustrated in Section 1 (Chapter 1), electron shuttling compounds or mediators are those which shuttle electrons from bacterial cell surface to anode. They are both exogenous, for example, neutral red, anthraquinone-2-6, disulfonate (AQDS), thionin, potassium ferricyanide, and methyl viologen, and endogenous, for example, pyocyanin and related compounds produced by Pseudomonas aeruginosa (Logan, 2008). As an effect of exogenous mediators/shuttle addition, enhancement in the generation of current and electricity is observed (Allen and Bennetto, 1993). The limiting step in MFC for dye removal is the transfer of electrons from a bacterial cell to azo dye and anodic electrode. The effect of the shuttle on dye removal using exogenous mediators was focused on to enhance dye degradation. Sun et al. (2013b) explained the phenomenon of dye removal and simultaneous electricity generation with the addition of synthetic and natural mediators. The MFC with 0.005 mM AQDS (anthraquinone-2,6-disulfonic disodium salt), 0.005 mM riboflavin (RF), or 1 g/L humic acid showed 36%, 26%, and 15% increase in maximum power density along with 394%, 450%, and 258% increases in decolorization rates of Congo red, respectively, when compared with mediator-free MFC. Various redox mediators such as flavin-based compounds such as RF, as well as quinone-based compounds such as anthraquinone-2,6-disulfonic acid disodium salt (AQDS), have been documented to accelerate the electron transfer; further accelerate the decolorization of azo dye and enhance electricity generation in MFC (Field and Brady, 2003). A small concentration of redox mediators is enough for this type of electron transfer.

4.3

Microbial diversity involved in the breakdown of dye in microbial fuel cell

The dye removal in MFCs greatly depends upon the anodic bacteria activity for the breakdown of azo bonds. Various microbial characterization studies done so far for MFCs, particularly involved in the degradation of dye and electricity generation, reported Pseudomonas putida, S. oneidensis, Azospirillum, Methylobacterium, Rhodobacter, Desulfovibrio, Trichococcus, Bacteroides, Pseudomonas delhiensis, Bacillus circulans, and Aquamicrobium defluvi strains in anode biofilm (Wang et al., 2014; Sun et al., 2013a, 2016; Xu et al., 2013; Fernando et al., 2012). In fact,

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the presence or absence of dye altered the microbial species in MFC. Geobacter sp., the well-known exoelectrogen, formed biofilm over the anode both in the presence and absence of Congo red, while other microbial communities such as Azospirillum, Methylobacterium, Rhodobacter, Desulfovibrio, Trichococcus, and Bacteroides sp. were present only in the presence of dye (Sun et al., 2013a). The consorted metabolic functioning of electricity generators and dye reducers is responsible for the better performance of MFC. The various isolates from the anode fed with different dye-containing wastewater are found to have a synergistic effect on each other (Wang et al., 2014; Fernando et al., 2012). Xu et al. (2013) and Sun et al. (2016) reported the supporting fact stating that two species, Pseudomonas and Bacillus, were prevalent in MFC anode treating dye wastewater of which Pseudomonas sp. is well-known exoelectrogens and Bacillus sp. is responsible for decolorization and degradation of azo dye. Hou et al. (2012) studied the effect of different cathode types on electricity generation and dye decolorization and revealed that the biocathode enhanced Congo red azo dye decolorization rate tremendously than that of air-cathode by altering its anodic biofilm strains. They also discovered that biocathode-manipulated anodic strains were Deltaproteobacteria, Firmicutes, Chlorobi, and Bacteroidetes, whereas the new phyla of Alphaproteobacteria and Betaproteobacteria were detected in air-cathode MFC.

4.4

Toxicity of treated dye wastewater

The treatment of dye wastewater is essential because of its toxicity on living beings. But, merely the treatment of dye wastewater does not necessarily mean detoxification. The nonspecific reduction of azo bonds in azo dyes, by an enzyme named azoreductase found in various microorganisms and in all tested mammals, including humans, under anaerobic conditions can sometimes end up to products with toxicity even higher than their parent compound (Weisburger, 2002). Methyl red, for instance, is mutagenic in nature, and its anaerobic degradation product N, N-dimethyl-phenylenediamine is aromatic amine, which is toxic and mutagenic as well (Wong and Yuen, 1998). Similarly, various other dye such as Alizarin Yellow R, MO, Acid Orange 7 (AO-7) on treatment with MFC undergo on anaerobic degradation and reduce to aromatic amines (Cui et al., 2012; Zhang et al., 2015; Thung et al., 2015; Fernando et al., 2014); many of which show very high-level toxicity, mutagenicity, and carcinogenicity (Weisburger, 2002; Fernando et al., 2014; Bafana et al., 2011). Therefore, it becomes very crucial to analyze the toxicity of the resultant compound after degradation by performing phytotoxicity and microbial toxicity tests (Ali, 2010) for large-scale applications. Since the complete reduction of azo bonds is only feasible in anaerobic conditions that result in aromatic amine production, the further aerobic treatment of degraded moieties completely mineralize and detoxify them (Cui et al., 2012; Abbassi et al., 2013; Kalathil et al., 2012; Fernando et al., 2014). The examination of toxicity of real dye wastewater and its treated effluent in a GACB-MFC through

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resazurin reduction test, an indicator of cell viability, showed 59% 6 9% and 7.9% 6 2% inhibition, respectively, indicating effluent is almost nontoxic. This was due to the high surface area of electrodes for electron production in the anode, and subsequent oxygen and dye reduction, as well as oxidation of degraded intermediates at the cathode (Kalathil et al., 2012). Fernando et al. (2014), while investigating the treatment performance of an integrated MFC-aerobic system for AO-7 dye-containing wastewater, conducted the bioluminescence-based toxicity test of samples at various stage. They found the half-maximal luminescence inhibition value (EC50) for the AO-7 containing synthetic wastewater influent, MFC effluent, and postaerobic treatment as 127, 63.3, and 637 mg COD/L, respectively, indicating toxicity increase from influent to MFC stage treatment and further decrease on subjection to postaerobic treatment. Toxicity finally reduced by 5-fold and 10-fold from influent and MFC stage effluent, respectively.

4.5

Microbial fuel cell
coupled techniques for textile wastewater treatment

The conventional biological processes hold the potential to treat azo dyes only up to a certain limit. To enhance the treatment efficiency, the bioelectrochemical treatment system took over the role, but even better was demanded while considering the ill-effects of discharging the effluent received on the ecological system (Sreelatha et al., 2016). Various MFC-coupled systems for the complete or partial degradation of the dye to form nontoxic products explored so far are MFCintegrated constructed wetlands (CWs), bioelectro-Fenton technology, MFC-biofilm electrode reactor coupled system, MFC-photoelectrocatalytic cell combined system, etc.

4.5.1 Microbial fuel cell
integrated constructed wetlands The application of CW-MFC technology has accelerated the biological wastewater treatment process and scaled up the bioelectrochemical treatment process to the field level (Srivastava et al., 2018a,b, 2020a,b; Yadav et al., 2018; Yadav, 2010). The coupling of MFC into CW is found to promote the activity of anaerobic electro-active bacteria (EAB) in the anodic strata of CW. This enhances the oxidation of cosubstrate/electron donor by EAB to produce readily available electrons that are accepted by azo dye as terminal electron acceptor and result in indirect and accelerated reductive cleavage of azo bonds (N 5 N) (Li et al., 2010; Yadav et al., 2012a,b; Fang et al., 2013). The wetland plants contribute significantly to the treatment process in two ways that are by harboring microbial assemblage in its rhizome and leaking out oxygen in the form of rhizospheric leakage. The thriving microbial assemblages feed on the root exudates for their metabolism and favor microbial oxidation of the azo dye’s reduced products that fasten their mineralization. The plants uptake

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some of the reduced and simplified products of dye, produced in the anaerobic region, for their growth. Thus, the synergistic approach among the two benefits of the dye removal process. Along with the high removal of dye, 80% and 91%, at HRT of 3 and 4 days, respectively, Yadav et al. (2012a,b) and Fang et al. (2013) have generated the maximum power density of 15.73 mW/m2 and 0.302 W/m3, respectively. The effect of varying HRT in the CW-MFC treatment system on the dye removal process and electricity generation was studied. It was found that HRT significantly affects the anodic dye degradation, but the linear relationship does not exist among the two. With increasing HRT, dye reduction increases up to a certain limit, but the further elongation of HRT shows a depreciating decolorization rate. Fang et al. (2015) reported the highest decolorization rate of 92.83% with power density of 0.0619 W/m3 at 3 days HRT, which decreased with subsequent elongation of HRT. The COD removal rate, power density, columbic efficiency, and open-circuit voltage followed the same trend in their study. Similarly, Fang et al. (2015) inferred that the COD of wastewater affected dye removal positively. With the increase in COD concentration, the decolorization rate also increases. Thus, the optimized operating conditions of CW-MFC show high decolorization performance and electricity generation.

4.5.2 Microbial fuel cell couple aerobic biocontact oxidation reactor system The azo dyes are typically recalcitrant to microbial aerobic oxidation. They undergo reductive degradation under anaerobic conditions to cleave at azo bond (N 5 N), forming daughter compounds that are colorless (Fernando et al., 2014, 2016; Kalathil et al., 2012; Cui et al., 2012). The results of various open and closed circuit comparative studies emphasized the contribution of the MFC stage alone in the decolorization rate (Fernando et al., 2014). However, the further degradation of toxic products to nontoxic simpler forms is feasible only in the aerobic condition (Cui et al., 2012; Fernando et al., 2014). The aerobically biotransformed simpler compounds are more accessible (bioavailable) to microbes for enhanced growth that further reduced the COD concentration. The removal efficiency of two azo dyes, AO-7 and Alizarin Yellow R in MFC-ABOR, are more than 90% (Fernando et al., 2014; Cui et al., 2012). The untreated dye fraction in MFC effluent remains recalcitrant even upon aerobic degradation and contributes toxicity. The operational parameters and electrode material need to be optimized to achieve maximum reduction of azo bonds. Fernando et al. (2014) in their comparative open and closed circuit study reported no adverse effect of high dye (AO-7) loading rate on the treatment performance, whereas Cui et al. (2012) stated inhibition in the metabolic activity of anode respiring bacteria upon high loading of Alizarin Yellow R. The toxicity test for AO-7 revealed that the aerobic posttreatment system reduced the toxicity up to 10-fold than the MFC effluent that abundantly consisted for amines, and 5-fold compared to the influent synthetic wastewater.

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4.5.3 Bioelectro-Fenton technology-microbial fuel cell The MFC is characterized by the probable in situ hydrogen peroxide (H2O2) production in the cathodic chamber based on the reduction of O2 molecule by electrons flowing from anodic to the cathodic chamber. H2O2 is formed as a result of a slowstep reaction (partial oxygen reduction with the consumption of only two electrons) during the two-step reduction process of oxygen to water (Rozendal et al., 2009). O2 1 2H1 1 2e2 ! H2 O2 Fe12 1 H2 O2 ! Fe13 1 OH 1 OH2 Earlier, the hydrogen peroxide production was considered as the futile reaction as it depreciates the current and power densities by inefficient utilization of electrons. Few researchers with the positive outlook considered the process utility by incorporating Fenton reaction into MFC for the in situ H2O2 oxidation to hydroxyl ions and radicals. The hydroxyl radicals are strong oxidizing agents with a redox potential of 22.8 eV that have been studied to oxidize recalcitrant organics such as azo dyes (Asghar et al., 2015). Zhuang et al. (2010) explored the MFC-integrated Fenton process (BioelectroFenton) for the decolorization of Rhodamine B using different cathodes and with different circuit connection patterns. The Fe@Fe2O3/NCF cathode performed best, giving 79.0% decolorization in closed circuit (1000 Ω) and maximum power density of 307 mW/m2, in 24 h retention time. The short-circuit connection enhanced the decolorization up to 95% owing to the enhanced current density in just 12 h. The findings of Fu et al. (2010) showed that the in situ generation of H2O2 for degradation of azo dye under neutral cathodic conditions could achieve 76.5% degradation efficiency in 1 h by utilizing 0.5 mmole/L Fe13 as a catalyst. The functional behavior of mixed culture was studied by De Dios et al. (2013) for decolorization of lissamine green B and crystal violet, simultaneously by the bioelectro-Fenton process. The synergy between the T. versicolor (fungi) and S. oneidensis MR-1 (bacteria) reduced the electron transport barriers from the bacteria to the anode, since the hyphae framework of fungi was utilized by bacteria for growth. Moreover, the symbiotic relation and catabolic cooperation enhanced the biodegradation efficiency for organics. The maximum volumetric power density recorded was of 0.78 W/m3, which is one of the highest achieved to date. Such synergy can further be used to support the external electrochemical processes.

4.5.4 Electrolysis cell combined with a microbial fuel cell (MFC-MEC) The performance of MFC for dye removal has been investigated thoroughly, but various aspects such as electrode material, type, and other modifications along with operational conditions and reactor configurations are still being studied to enhance the decolorization performance. The electrochemical reduction of recalcitrant dyes at cathode needs to overcome a specific lower potential that requires

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additional voltage than the one produced by MFC for enhanced dye and other pollutants removal (Srivastava et al., 2018a,b, 2019a, 2020a,b; Li et al., 2016; Wang et al., 2013). The additional supply of voltage provides a good alternative to MFC alone. As an advantage, the performance enhances as well as the reduced external power supply is needed as it complements the MFC voltage to match up the activation voltage needed for reduction reaction occur. Wang et al. (2013) have used an external voltage of 0.5 V between the anode and biocathode to obtain the enhanced dye (Amido black 10B) removal up to 81.7% in 10 h of HRT. The anode and cathode potentials were 20.5 and 21.0 V (vs saturated calomel electrode), respectively. Li et al. (2016) also studied the microbial electrolysis cell incorporated microbial fuel cell, and the results revealed that in a coupled system, the decolorization rate was 75.2% that is an improvement over to MFC alone, which produces 36.5%.

4.6

Research gap

The literature concerning dye removal and concomitant electricity production in an MFC system is still in its infancy and needs much more attention for its future application. The studies conducted so far related to design and basic necessary operating parameters for decolorization and detoxification have illustrated the performances achieved, but very less effort has been made to enhance the achieved treatment performance of MFC with modification in the anode (bioanode), cathode (biocathode), and membrane (Srivastava et al., 2017, 2019b; Kalathil et al., 2011, 2012; Bakhshian et al., 2011). The supporting data for any stated fact are severely lacking and provide an opportunity for duplication and/or conductance of studies in a similar direction. The microbial communities in MFC play a vital role in dye degradation and electricity generation. The phylogeny and strains of bacterium involved in the electricity generation during dye removal have been well illustrated in literature, but the dye removing strain flourishing symbiotically in the anodic biofilm, has neither been well characterized, nor the biochemistry of this is explored. The treatment of toxic dye wastewater in MFC involves the breakdown of azo bonds that forms aromatic amines that are toxic as well. Often, higher toxicity of intermediate products is reported than the parent dye. The complete mineralization and detoxification of dye in wastewater requires intensive oxidation reactions that may or may not be feasible in all MFC systems. The incorporation of MFC with other supplementary oxidative techniques (hybrid systems) needs to be investigated for large-scale application of MFC as dye detoxifiers, rather than MFC as a standalone device. Also, the decolorization of dye, dye removal, and less-toxic intermediates obtained in MFC effluent does not mitigate the problem of dye-containing wastewater. The toxicity test for treated effluent needs to be performed before discharging it. Almost no literature accomplishes the toxicity test for MFC-treated dye wastewater to date.

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Acknowledgments YM gratefully acknowledges the GATE-JRF fellowship from CSIR, PT thankfully acknowledges the JRF fellowship from UGC, SG, and AKY acknowledges the funding from NASF (IARI).

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571. Sreelatha, S., Velvizhi, G., Naresh Kumar, A., Venkata Mohan, S., 2016. Functional behavior of bio-electrochemical treatment system with increasing azo dye concentrations: synergistic interactions of biocatalyst and electrode assembly. Bioresour. Technol. 213, 11
20. Stams, A.J.M., de Bok, F.A.M., Plugge, C.M., van Eekert, M.H.A., Dolfing, J., Schraa, G., 2006. Exocellular electron transfer inanaerobic microbial communities. Environ. Microbiol. 8 (3), 371e382. Sun, J., Hu, Y.Y., Bi, Z., Cao, Y.Q., 2009a. Simultaneous decolorization of azo dye and bioelectricity generation using a microfiltration membrane air-cathode single-chamber microbial fuel cell. Bioresour. Technol. 100 (13), 3185
3192. Sun, J., Hu, Y., Bi, Z., Cao, Y., 2009b. Improved performance of air-cathode single-chamber microbial fuel cell for wastewater treatment using microfiltration membranes and multiple sludge inoculation. J. Power Sources 187 (2), 471
479. Sun, J., Bi, Z., Hou, B., Cao, Y.Q., Hu, Y.Y., 2011. Further treatment of decolorization liquid of azo dye coupled with increased power production using microbial fuel cell equipped with an aerobic biocathode. Water Res. 45 (1), 283
291. Sun, J., Li, W., Li, Y., Hu, Y., Zhang, Y., 2013b. Redox mediator enhanced simultaneous decolorization of azo dye and bioelectricity generation in air-cathode microbial fuel cell. Bioresour. Technol. 142, 407
414. Sun, J., Li, Y., Hu, Y., Hou, B., Zhang, Y., Li, S., 2013a. Understanding the degradation of Congo red and bacterial diversity in an air
cathode microbial fuel cell being evaluated for simultaneous azo dye removal from wastewater and bioelectricity generation. Appl. Microbiol. Biotechnol. 97 (8), 3711
3719. Sun, J., Cai, B., Zhang, Y., Peng, Y., Chang, K., Ning, X., et al., 2016. Regulation of biocathode microbial fuel cell performance with respect to azo dye degradation and electricity generation via the selection of anodic inoculum. Int. J. Hydrogen Energy 41 (9), 5141
5150. Tartakovsky, B., Guiot, S.R., 2006. A comparison of air and hydrogen peroxide oxygenated microbial fuel cell reactors. Biotechnol. Progr. 22, 241
246. TerHeijne, A., Hamelers, H.V.M., Buisman, C.J.N., 2007. Microbialfuel cell operation with continuous biological ferrous ironoxidation of the catholyte. Environ. Sci. Technol. 41 (11), 4130e4134. Thung, W.E., Ong, S.A., Ho, L.N., Wong, Y.S., Ridwan, F., Oon, Y.L., et al., 2015. A highly efficient single chambered up-flow membrane-less microbial fuel cell for treatment of azo dye Acid Orange 7-containing wastewater. Bioresour. Technol. 197, 284
288. Van der Zee, F.P., Lettinga, G., Field, J.A., 2001. Azo dye decolourisation by anaerobic granular sludge. Chemosphere 44 (5), 1169
1176.

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Varanasi, J.L., Das, D., 2018. Bioremediation and power generation from organic wastes using microbial fuel cell. Microbial Fuel Cell. Springer, Cham, pp. 285
306. Wang, Y.Z., Wang, A.J., Liu, W.Z., Kong, D.Y., Tan, W.B., Liu, C., 2013. Accelerated azo dye removal by biocathode formation in single-chamber biocatalyzed electrolysis systems. Bioresour. Technol. 146, 740
743. Wang, V.B., Chua, S.L., Cai, Z., Sivakumar, K., Zhang, Q., Kjelleberg, S., et al., 2014. A stable synergistic microbial consortium for simultaneous azo dye removal and bioelectricity generation. Bioresour. Technol. 155, 71
76. Weisburger, J.H., 2002. Comments on the history and importance of aromatic and heterocyclic amines in public health. Mutat. Res./Fundam. Mol. Mech. Mutagen. 506, 9
20. Wong, P.K., Yuen, P.Y., 1998. Decolourization and biodegradation of N, N0 -dimethyl-p-phenylenediamine by Klebsiella pneumoniae RS-13 and Acetobacter liquefaciens S-1. J. Appl. Microbiol. 85 (1), 79
87. Xu, Q., Sun, J., Hu, Y.Y., Chen, J., Li, W.J., 2013. Characterization and interactions of anodic isolates in microbial fuel cells explored for simultaneous electricity generation and Congo red decolorization. Bioresour. Technol. 142, 101
108. Yadav, A.K., 2010. Design and development of novel constructed wetland cum microbial fuel cell for electricity production and wastewater treatment. In: Proc. of the 12th IWA International Conference on Wetland Systems for Water Pollution Control. Venice, pp. 1085
1089. Yadav, A., Dash, P., Mohanty, A., Abbassi, R., Mishra, B.K., 2012a. Performance assessment of innovative constructed wetland-microbial fuel cell for electricity production and dye removal. Ecol. Eng. 47, 126
131. Yadav, A.K., Jena, S., Acharya, B.C., Mishra, B.K., 2012b. Removal of azo dye in innovative constructed wetlands: influence of iron scrap and sulfate reducing bacterial enrichment. Ecol. Eng. 49, 53
58. Yadav, A.K., Srivastava, P., Kumar, N., Abbassi, R., Mishra, B.K., 2018. Constructed wetland microbial fuel cell: an emerging integrated technology for potential industrial wastewater treatment and bioelectricity generation. In: Stefanakis, A.I. (Ed.), Constructed Wetlands for Industrial Wastewater Treatment. John Wiley & Sons, Hoboken, NJ, pp. 493
510. , 2018. You, S.J., Zhao, Q.L., Zhang, J.N., Jiang, J.Q., Zhao, S.Q., 2006. A microbial fuel cell using permanganate as the cathodic electron acceptor. J. Power Sources 162 (2), 1409
1415. Zhang, L., Liu, C., Zhuang, L., Li, W., Zhou, S., Zhang, J., 2009. Manganese dioxide as an alternative cathodic catalyst to platinum in microbial fuel cells. Biosens. Bioelectron. 24 (9), 2825
2829. Zhang, B., Wang, Z., Zhou, X., Shi, C., Guo, H., Feng, C., 2015. Electrochemical decolorization of methyl orange powered by bioelectricity from single-chamber microbial fuel cells. Bioresour. Technol. 181, 360
362. Zhuang, L., Zhou, S., Yuan, Y., Liu, M., Wang, Y., 2010. A novel bioelectro-Fenton system for coupling anodic COD removal with cathodic dye degradation. Chem. Eng. J. 163 (1), 160
163.

Further reading Bond, D.R., Lovley, D.R., 2003. Electricity production by Geobacter sulfurreducens attached to electrodes. Appl. Environ. Microbiol. 69 (3), 1548
1555.

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Chaudhuri, S.K., Lovley, D.R., 2003. Electricity generation by direct oxidation of glucose in mediatorless microbial fuel cells. Nat. Biotechnol. 21 (10), 1229
1232. Choudhury, P., Uday, U.S.P., Mahata, N., Tiwari, O.N., Ray, R.N., Bandyopadhyay, T.K., et al., 2017. Performance improvement of microbial fuel cells for waste water treatment along with value addition: a review on past achievements and recent perspectives. Renew. Sustain. Energy Rev. 79, 372
389. Lovley, D.R., 2008. The microbe electric: conversion of organic matter to electricity. Curr. Opin. Biotechnol. 19 (6), 564
571. Malvankar, N.S., Lovley, D.R., 2014. Microbial nanowires for bioenergy applications. Curr. Opin. Biotechnol. 27, 88
95.

5

Agro-industrial wastewater treatment in microbial fuel cells Silvia Bolognesi, Daniele Cecconet and Andrea G. Capodaglio Department of Civil Engineering and Architecture, University of Pavia, Pavia, Italy

Chapter Outline 5.1 5.2 5.3 5.4

Introduction 93 Use of agro-industrial wastewater as substrate for microbial fuel cells Dairy industry wastewater 96 Brewery and winery industry 101

95

5.4.1 Brewery wastewater 101 5.4.2 Winery wastewater 102

5.5 Agro-industrial wastewaters and by-products 107 5.5.1 Palm oil industry wastewater 107 5.5.2 Agricultural products processing wastewater 109 5.5.3 Agricultural residues 113

5.6 Livestock industry wastewater 117 5.7 Challenges in using microbial fuel cells 5.8 Conclusion 123 References 123

5.1

120

Introduction

The growing pressure exerted from massive demographic increase on society and resources, in combination with challenges determined by the related energy needs and the stricter rules on disposal of wastes, has recently imposed a marked turnaround in the definition of national and international environmental goals. To face increasing environmental issues, balance must be reached between the development of clean and renewable energy sources, waste reduction, and recovery of reusable materials. Microbial fuel cells (MFCs) is a technology that may fulfill part of these needs, allowing electricity to be recovered directly from an organic waste matrix, in concomitance with pollutants removal. MFCs can be defined as bioelectrochemical systems (BESs) that directly convert the energy accumulated in the chemical bonds of organic matter into electrical energy, by means of electrochemically active bacteria (EABs), acting as catalyzers of the reactions occurring within the process (Capodaglio et al., 2013). In an MFC, bacteria oxidize the organic substrate, producing electrons traveling through a series of respiratory enzymes within organic cells. Electrons are then released to a terminal electron acceptor (TEA, typically, oxygen) Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00005-9 © 2020 Elsevier Inc. All rights reserved.

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that is consequently reduced. MFCs technical feasibility has been supported by the contribution of many researchers, who, from the early 2000s, have analyzed their operating principles, configurations, materials, methods and measuring instruments, promoting their technological evolution and an expansion of the application domain. Among the reasons for the great enthusiasm that MFCs have aroused in the scientific community is the urgency of developing solutions to face two of the great environmental emergencies of today: availability of clean and safe water, and the need for energy; in this context, MFCs are a promising technology to achieve both wastewater treatment and energy production (ElMekawy et al., 2015). Application of MFCs for organic matter removal has provided excellent results: in several cases, average COD (chemical oxygen demand) removal efficiencies up and above 90% have been achieved. Although studies reported in the literature confirm the promising nature of this technology as a renewable energy source, MFCs still require further improvements that make them economically attractive in this sense. Despite an organic matter removal rate comparable to that of conventional (e.g., activated sludge) treatment systems [up to 7 kg CODREMOVED/(m3 day), while ranges of 0.5 2 kg CODREMOVED/(m3 day) are reported for activated sludge and 8 20 kg CODREMOVED/(m3 day) for anaerobic processes, respectively] (Logan et al., 2006; Molognoni et al., 2017), power and current produced are often insufficient to justify a practical implementation of MFCs on an industrial scale solely for that purpose. Compared to anaerobic digestion, for example, the gain from electricity production in MFCs is too low to offset the high investment and operational costs (Rozendal et al., 2008). Fig. 5.1 represents a schematic of possible wastewater treated with MFC and TEAs required for the redox reaction.

Figure 5.1 Schematic of chemical reactions from oxidation of various wastewaters, and terminal electron acceptors, in an MFC. MFC, Microbial fuel cell.

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95

Agro-industries convert raw agricultural materials into value-added products while generating income and employment and contributing to overall economic development in both developed and developing countries. Food processing converts relatively bulky, perishable, and typically inedible raw materials into more useful, shelf-stable, and palatable foods or potable beverages. Processing contributes to food security by minimizing waste and loss in the food chain and by increasing food availability and marketability, and to improve its quality and safety (Food and Agriculture Organization, n.d.). Agro-industrial wastewater is characterized by high organic strength that may induce adverse effects in water bodies (Alayu and Yirgu, 2018). Agro-industrial expansion in both developed and developing countries are major contributors of environmental pollution (Rajagopal et al., 2013). Treatment technology applied to agro-industrial wastewaters depend upon agroindustrial wastewater characteristics (and on the processes originating them) and the required effluent quality levels, considering also cost, operation, and efficiency of the processes. Effluent quality control is aimed to public health and environmental protection, with limits set depending on the final use of water (i.e., recreation, irrigation, and/or water supply) area of discharge (i.e., eutrophication-sensitive, or not). MFCs have been tested for treatment of many industrial wastewater types so far, with the purpose of finding the most convenient solution (in terms of energy recovery and investment costs) for upscaling of the technology. The use of wastewater from the agro-food industry seems particularly promising, due to its high organic matter content (COD .1 g/L) and its easy biodegradability, in terms of biological oxygen demand (BOD), presenting generally a BOD/COD ratio .60% (Callegari et al., 2018; Capodaglio et al., 2016; Cercado-quezada et al., 2010). Herein, in this chapter, an overall view of the state-of-the-art and of several experiences concerning MFC treating agro-industry wastewaters is reported.

5.2

Use of agro-industrial wastewater as substrate for microbial fuel cells

Agro-industry, and in particular the food industry, generates large amounts of liquid, solid, and gaseous wastes. In general, agro-food processing wastes are composed of large amounts of organic materials (carbohydrate, protein, fat, oil, etc.), with high values of BOD, COD, and suspended solids. They also have a high potential to cause severe pollution problems, if not properly managed or treated (Prasertsan et al., 2009; Kim et al., 2013), as they are characterized by the presence of total solids (TS), total nitrogen, total phosphorus (TP), BOD and COD, and pathogens (Alayu and Yirgu, 2018). Water pollution seems to be the most serious concern in many agro-industries when dealing with waste, since solid wastes have a much higher opportunity for recovery or utilization. Water is commonly used in food processing as a main ingredient (e.g., in beverage industries), but also for washing and processing raw materials, cooling, and cleaning. The volume of water

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used, and the types and concentration of pollutants are extremely variable, even within the same sector. Generally, biological processes (aerobic or anaerobic) are widely used for the treatment of agro-industrial wastewaters (Satyanarayan et al., 2005). Researchers have investigated treatment of a wide range of wastewaters by different MFC designs and operational conditions, reporting results in term of different parameters such as coulombic efficiency (CE), COD removal efficiency (ηCOD), and electrical energy production in terms of current density or power density (Molognoni et al., 2014). Substrate composition is considered as one of most important factors affecting electricity generation in MFC (Pant et al., 2010). Agro-food substrates seem particularly promising for MFCs application and scaling-up. However, complex substrates increase bacterial community complexity and can lead to interrelated connections between different microbial species present (Vilajeliu-Pons et al., 2016). Anodic chamber anaerobic conditions may lead to the appearance of unwanted side-reactions, such as methanogenesis or heterotrophic denitrification (Capodaglio et al., 2015, 2017). High fermented substrate concentrations favor methanogenic activity compared with exo-electrogenesis, reducing MFCs’ CE (Pinto et al., 2010). In the chapter, characteristics of substrates used and experiences from many authors treating dairy, brewery and winery, agro-industrial, and livestock industry with MFCs are reported.

5.3

Dairy industry wastewater

Dairy industry involves processing raw milk into several products, during which, it generates high-strength wastewaters characterized by high BOD and COD concentrations (Demirel et al., 2005), which incidentally make them a chemical energy rich substrate, and therefore a valuable, alternative energy source (Najafpour et al., 2008). In the dairy products market a prominent position is occupied by Italy, contributing approximately 10% of the European cow milk production; in turn, the European cow milk production covers more than 22% of the world’s production. Milk from other dairy species should also be added to this productive wealth, such as sheep milk (660.453 t, or 5.4% of the total), goat milk (120.790 t equal to 1% of the total) and buffalo milk (150.500 t equal to 1.2% of the total) having considerable relevance in the Italian production context. Considering cheese production, Italy is ranked fifth with 6% of world’s production, after United States, Russia, France, and Germany (Laraia et al., 2001). According to the Italian Environmental Agency (APAT), estimated national dairy industry wastewater production is about 19 million tons per year. At the basis of the appeal for valorization of dairy wastewater (DWW), there is the huge amount of water implied in this productive process, approximately equivalent to a city of 200,000 inhabitants in terms of flow alone. Dairy industry is one of the most water-consuming industry: nearly 2.5 times the

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97

Table 5.1 Chemical physical characterization of wastewater produced in a dairy industry Parameter

U.M.

Concentration range

TSS COD Biological oxygen demand (BOD5) Organic nitrogen (Norg) Ammonium (NH4-N) Nitrate (NO3-N) Total nitrogen (Ntot) Phosphorus (P) Chlorides (Cl) pH Alkalinity (HCO3)

mg SST/L mg O2/L mg O2/L mg N/L mg N/L mg N/L mg N/L mg P2O5/L mg Cl/L

250 2700 650 3000 300 1400 10 140 10 20 10 20 30 180 10 132 50 500 4 12 257 657

mg HCO3/L

TSS, Total suspended solids. Source: Based on APAT, 2007; Gutierrez et al., 1991; Passeggi et al., 2009.

volume of the milk processed is generated as wastewater, in the form of spent wash or spillage (Venkata Mohan et al., 2010). The volume, concentration, and composition of the waste deriving from dairy industry depend on the type of product (strongly diversified as milk, butter, yogurt, ice cream, various types of desserts and cheese), the processing line used, and the specific size of the factory (APAT, 2007). DWW characteristics include the presence of complex organics, such as polysaccharides, proteins, and lipids, which during hydrolysis are transformed into sugars, acids, and fatty acids. Sugars contribute to about 97% of the total COD of these streams. Biodegradable organics and nutrients could be beneficially used for renewable energy generation (Venkata Mohan et al., 2008). Chemical physical characterization of dairy industry wastewater is reported in Table 5.1. Due to its properties, DWW can be regarded as an efficient anolyte in MFCs (He et al., 2017). Considerable interest has been raised in the last decade on using MFC technology to treat dairy wastes (Cecconet et al., 2017, 2018a,b; Mahdi Mardanpour et al., 2012; Nimje et al., 2012). Cheese whey (CW), the liquid that remains after the precipitation and removal of milk casein during cheese-making, is a protein-rich by-product of the dairy industry. It contains highly biodegradable organic matter (Demirer et al., 2000), with chemical composition depending on quality and composition of milk and production techniques (i.e., amount of yeast and acid used for fermentation and coagulation, respectively) (Alayu and Yirgu, 2018). CW contains high-strength organic pollutants: COD and BOD5 values range from 60,000 100,000 to 40,000 60,000 mg/L, respectively (Chatzipaschali and Stamatis, 2012), and represent about 85% 95% of the original milk volume. CW contains nutrients, such as 4% 5% carbohydrates, proteins up to 1%, fats at about 0.4% 0.5%, and salts 1% 3% (Pandey et al., 2016) and retains 55% of milk nutrients. CW salts comprise net anodic compartment (NaCl) and KCl, calcium salts (primarily phosphate), and other components, such as lactic (less than 1%) and citric acids, nonprotein nitrogen

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compounds (urea and uric acid), and B group vitamins. CW disposal constitutes a serious environmental problem because of the high volumes produced and its high organic matter contents. Two main whey varieties are produced: acid (pH , 5) and sweet (pH 6 7) whey, according to the procedure used for casein precipitation (Gonza´lez Siso, 1996). Bioelectrochemical recovery of electricity from DWW and CW with MFCs has been investigated in recent years. Single-chamber, dual-chamber, and tubular MFCs have been tested, with electrodes of different materials, including carbon, graphite, stainless steel, or composites. Table 5.2 summarizes different experiences with MFCs using dairy industry wastewater and CW as anodic substrate. The highest reported power density (27 W/m3) was achieved by Cecconet et al. (2017), treating raw DWW using a two chamber MFC with granular graphite electrodes and use of a biocathode. The MFC was operated in continuous mode for 72 days, obtaining a maximum CE of 36% and up to 90% COD removal. Mahdi Mardanpour et al. (2012) also achieved high power density (20.2 W/m3) and the highest COD removal efficiency reported so far (91%) using a tubular MFC equipped with a 0.5 mg Pt/cm2 catalyst load on the cathode. The feed solution was replaced when the measured voltage between electrodes dropped below 20 mV, completing one cycle of operation. Other studies demonstrated that the use of cathode catalysts may be avoided, without negative effects on organic matter removal efficiency, but with considerable reduction of power density (90% lower) and CE (44% lower) (Elakkiya and Matheswaran, 2013; Venkata Mohan et al., 2010). Mansoorian (2016) achieved the highest CE (37.2%) using a catalyst-free and mediator-less MFC. Experiences of continuously fed MFCs treating DWWs achieved generally worse results than batch-fed systems, in terms of both power density and COD removal, but scored higher CEs (Faria et al., 2017b), with some exceptions (Cecconet et al., 2017). Venkata Mohan et al. (2010) observed increasing MFC electrical performance with increasing organic loading rates (OLRs) to the anode chamber. On the other hand, Elakkiya and Matheswaran (2013) noticed that high anolyte’s COD concentration (up to 2800 mg/L) can cause ionic exchange membrane fouling, and gradual decrease of electric production. Also, the development of unsustainably thick anodic biofilm, which will suffer from high concentration overpotentials, could result from high loads of proteins and lipids (Cercadoquezada et al., 2010). CW has been used as a substrate by Antonopoulou et al. (2010) in a batch-fed, two-chambered H-type MFC with carbon paper anode electrode and Pt-coated (0.5 mg/cm2) cathode, achieving a maximum power density of 18.4 mW/m2. Tremouli et al. (2013) achieved a very high COD removal efficiency (94%) using a dual-chamber fed-batch MFC, and 11.3% CE, confirming CW as a suitable substrate for MFCs. Kelly and He (2014) operated four fed-batch mode cylindrical MFCs with carbon brushes as electrodes and carbon powder as cathodic catalyst. Each MFC treated a different cheese production waste as anodic substrate; the one fed with CW achieved maximum energy production (1.18 kWh/m3).

Table 5.2 List of dairy industry wastewater used as substrate in microbial fuel cells (MFCs) and their respective performances. Substrate

MFC type

Feeding mode

Volume (NAC) (mL)

Anode material

Cathode material

Substrate concentration

COD removal (ηCOD)

Power density (max)a

CE

Ref.

Dairy industry wastewater

Annular singlechamber spiral MFC

Batch

90

Graphite coated stainlesssteel mesh

1000 mg COD/L

91%

20.2 W/m3

26.9%

Mahdi Mardanpour et al. (2012)

Dairy industry wastewater

Continuous

2000

Plain graphite plate (188 cm2)

53.22 kg COD/m3

90.5%

621.1 mW/m2

37.2%

Mansoorian (2016)

Dairy industry wastewater

Catalyst-less and mediatorless membrane MFC Two-chamber MFC

Carbon cloth type B (30% wet proofing), 0.5 mg/cm2 Pt loading Plain graphite plate (188 cm2)

Semicontinuous

430

Granular graphite (800 g)

1500 mg COD/L

85%

27 W/m3

36%

Cecconet et al. (2017)

Dairy industry wastewater

Two-chamber MFC

Semicontinuous

430

Granular graphite (800 g)

1884 mg COD/L

71.81% 6 14.53%

15.40 6 1.89 W/ m3

11.94% 6 6.94%

Cecconet et al. (2018a,b)

Dairy industry wastewater

Dualchambered MFC

b

123

Granular graphite (800 g) Granular graphite (800 g) Carbon fiber brush electrode

3400 mg COD/L

82% 86%

3.2 W/m3

b

Cetinkaya et al. (2015)

Dairy industry wastewater

Dualchambered MFC Dualchambered MFC Dualchambered MFC

Continuous

350

Carbon sheet (72 cm2)

Carbon fiber brush electrode PtxNiy/C bimetallic electrocatalyst Carbon sheet (72 cm2)

500 6 100 mg COD/ L

63%

1.9 W/m3

24%

Faria et al. (2017a)

Batch

300

Plain graphite plate (50 cm2)

3200 mg COD/L

91%

2.7 W/m3

17%

Batch

480 550

Plain graphite plate (50 cm2) Plain graphite plate (70 cm2)

Plain graphite plate (70 cm2)

3700 mg COD/L

95%

1.1 W/m3

14%

Elakkiya and Matheswaran (2013) Venkata Mohan et al. (2010)

Dairy industry wastewater Dairy industry wastewater

(Continued)

Table 5.2 (Continued) Substrate

MFC type

Feeding mode

Volume (NAC) (mL)

Anode material

Cathode material

Substrate concentration

COD removal (ηCOD)

Power density (max)a

CE

Ref.

Cheese whey (diluted)

Dualchambered tubular MFC

Batch/ continuous

500

20 cm carbon fiber brush

1134 mg COD/L

74.8%

1.3 6 0.5 W/m3

b

Kelly and He (2014)

Cheese whey

Dualchambered MFC

Batch

310

0.73 g COD/L

b

18.4 mW/m2

1.9%

Antonopoulou et al. (2010)

Cheese whey (filter sterilized)

Dualchambered MFC

Batch

310

Teflon treated carbon fiber paper 10 wt. % wet proofing Plain carbon paper 10 wt. % wet proofing

Carbon cloth with activated carbon powder (9 mg/cm2) as catalyst Carbon cloth containing 0.5 mg/cm2 of Pt catalyst

6.7 g COD/L

94%

46 mW/m2

11.3%

Tremouli et al. (2013)

COD, Chemical oxygen demand. a If not otherwise specified, power density is normalized to anodic chamber volume (W/m3) or anode electrode surface (mW/m2). b Data not available.

Carbon cloth containing 0.5 mg/cm2 of Pt catalyst

Agro-industrial wastewater treatment in microbial fuel cells

5.4

101

Brewery and winery industry

5.4.1 Brewery wastewater Brewery wastewater (BW) is an agro-industrial waste generated in large quantities during beer production, by many different unit operations (saccharification, fermentation, cooling, washing, etc.) (Lu et al., 2017). According to Food and Agriculture Organization (n.d.) statistics, beer production from barley in 2014 reached 230 million tonnes worldwide. Considering only the European production, 43 million tonnes of beer have produced, with top production occurring in Germany (8.7 million tonnes), United Kingdom (4.1 million tonnes), Poland (4 million tonnes), and Spain (3.3 million tonnes). Water contributes to around 95% of the ingredients of beer and, for this reason, many breweries derive it from artesian wells, and are located in areas with ample water tables or rivers (Arantes et al., 2017). This use of hydro resources is, however, a concern, since an average of 4.5 L water per liter of produced beer is consumed, with peaks sometimes reaching up to 10 L/L, including the brewing, rinsing, and cooling processes (Simate et al., 2011). Water consumption is generally divided into two-third process water and one-third for cleaning operations (Fillaudeau et al., 2006). The brewery effluent quality depends on the different processes that take place within the brewery, particularly in the raw material handling, wort preparation, fermentation, filtration, and packaging (Alayu and Yirgu, 2018), that generate wastes containing residual amounts of raw materials, including solids, sugars, and yeasts, among other components. Filtration, equipment discharges and washing of containers, cleaning of tanks, vats, pipes, and floors all lead to the generation of large volumes of wastewater with high organic matter and suspended solids loads (Arantes et al., 2017). Table 5.3 shows the physical chemical characterization of BW. COD may vary from 2000 to 32,500 mg/L, with solids content following a similar trend, since a greater presence of solids results in higher COD. BW typically has a high COD from sugars, soluble starch, ethanol, volatile fatty acids, etc. (Goldammer, 2008), that are generally easily biodegradable. Table 5.3 Chemical physical characterization of brewery wastewater. Parameter

U.M.

Concentration range

TSS COD Biological oxygen demand (BOD5) Ammonium (NH4-N) Total Nitrogen (Ntot) Phosphorus (P) pH Alkalinity (HCO3)

mg SST/L mg O2/L mg O2/L mg N/L mg N/L mg P2O5/L

200 3000 2000 32,500 1200 3600 5 21.6 25 450 0.5 216 3 12 190 3173

mg HCO3/L

COD, Chemical oxygen demand; TSS, total suspended solids. Source: Adapted from Arantes, M.K., Alves, H.J., Sequinel, R., da Silva, E.A., 2017. Treatment of brewery wastewater and its use for biological production of methane and hydrogen. Int. J. Hydrogen Energy, 42(42), 26243 26256. https://doi.org/10.1016/j.ijhydene.2017.08.206.

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Nitrogen and phosphorus levels are mainly dependent on the handling of raw material and the amount of yeast present in the effluent (Simate et al., 2011). One reason for the large difference between the minimum and maximum values for COD, nitrogen, and phosphorus is related to the degree of dilution of the wastewater as volume of washing water varies. Overall, the parameters reported for BW indicate the possibility to treat these strong wastewaters by biological treatment, due to high organic matter and nitrogen presence. Conventional biological methods such as anaerobic reactors can be effective for BW treatment but require long treatment times (15 40 days) (Kavitha et al., 2013) if compared to aerobic solutions (8 6 12 h) (Nielsen et al., 2011). Feng et al. (2008) and Wen et al. (2010a) demonstrated the feasibility of using MFCs for the treatment of BW (Feng et al., 2008; Wen et al., 2010b). They treated BW with influent COD of 2250 6 418 mg/L, obtaining 87% COD removal efficiency and 38% CE. The authors observed that the efficiency of treatment process in terms of COD removal was not significantly influenced by temperature and conductivity but was greatly dependent on COD influent concentration, where higher strength wastewater showed higher COD removal rates. Wen et al. (2010a) draw similar conclusions for a small-scale MFC treating continuously fed BW. Lu et al. (2017) tested a 20 L MFC two-stage system fed with diluted BW for over a year. The 1-year operation period was divided into eight different periods (A H) under different operational conditions (HRT, hydraulic residence time, and applied external resistor). Period A (HRT 5 313 h, Rext510 Ω for both MFCs) and B (HRT 5 313 h, Rext53.7 Ω, and Rext55.3 Ω for the first and second stage, respectively) showed better results in terms of COD removal (up to 94.6%), while maximum power density was quite low (1.61 mW/m2 vs electrode surface). Some microbial community damage was reflected in higher flow rates needed after day 275 (period H). Dong et al. (2015) built and operated a continuously fed, stackable, and baffled 90 L MFC system for the treatment of raw BW. The system consisted of five stackable modules operated in self-sufficient manner for 6 months. The system achieved 82.7% COD removal and 8% CE. Overall energy production was 0.097 kWh/m3, sufficient to cover the energetic need required by the pumps. A schematic diagram of this system is represented in Fig. 5.2. A summary of experiences of MFCs treating BWs from several authors is reported in Table 5.5.

5.4.2 Winery wastewater Worldwide wine production is estimated to be around 250 million hectoliters per year. Wine production is one of the leading agro-food industries in Europe (about 62% of total amount of production worldwide), but recently it has also attained considerable importance in other parts of the world, such as Australia, Chile, the United States, South Africa, and China, with increasing influence on the economy of these countries (Calheiros et al., 2018; Masi et al., 2015).

Agro-industrial wastewater treatment in microbial fuel cells

103

Figure 5.2 Schematic diagram of the 90 L stackable baffled microbial fuel cell (Dong et al., 2015).

Winery wastewater (WW) production is not regular year round, its variability being caused by different factors, such as the adopted industrial process chain, its seasonality, and the kind of produced wine. Wastewater production varies in terms of duration, quantity, and composition, with peak at the harvest time (Calheiros et al., 2018). During the grape-processing period (vintage and racking), relevant fluxes of wastewater are generated and in the following months, when bottling and cleaning of containers are almost continuous operations, a massive amount of water is used (Masi et al., 2015). Winery industry is one of the most water-consuming industries: considering all the previously mentioned operations, a ratio of wastewater/wine of 14 L/0.5 L was reported (Oliveira and Duarte, 2010). WW is characterized by high organic loadings, up to 5 kg COD/(m3 day), mostly consisting of highly soluble sugars, alcohols, acids and recalcitrant compounds (e.g., polyphenols), tannins, and lignin. Ethanol and sugars (fructose and glucose) represent more than 90% of the organic load (Masi et al., 2015). In addition, WW is characterized by low pH and low concentrations of nitrogen and phosphorous, beside other factors that lead to caution in its discharge in the environment (Arienzo et al., 2009; Calheiros et al., 2018). Wineries often use wastewater for irrigation; otherwise, effluents may be discharged into surface water bodies or sewage system. When properly planned and controlled, the use of winery effluents for irrigation may be useful to agriculture contributing to save water and recycle nutrients (Oliveira and Duarte, 2010), while when discharging into surface water bodies, possible effects due to wastewater toxicity in certain periods of production have to be taken into account (Calheiros et al., 2018). Treatment and disposal of WWs is therefore one of the main environmental problems in wine-making industries. Chemical physical characterization of WW is reported in Table 5.4. In the last two decades, new methods have been proposed for the treatment of WW, consisting of both biological and electrochemical technologies. Their use is effective for removing organic matter, but they require high energy inputs, which increase the cost associated with the production of wine. In recent years, MFCs have attracted researchers’ attention and have been applied for WW treatment.

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Integrated Microbial Fuel Cells for Wastewater Treatment

Table 5.4 Chemical physical characterization of winery wastewater Parameter

U.M.

Concentration range

TSS COD Biological oxygen demand (BOD5) Total nitrogen (Ntot) Phosphorus (P) Chlorides (Cl) pH

mg SST/L mg O2/L mg O2/L mg N/L mg Ptot/L mg Cl/L

190 18,000 340 49,103 130 22,418 12 364 5.5 68 0 39.9 3.5 7.9

COD, Chemical oxygen demand; TSS, total suspended solids. Source: Based on Brito et al., 2004; Masi et al., 2015; Penteado et al., 2016.

Penteado et al. (2016) operated a two-chambered MFC fed with WW for 41 days, adding P and N to compensate their lack in the raw wastewater. This addition increased the energy production (from 105 to 465 mW/m2) and CE (from 2% to 15%), but COD removal efficiency remained low (17%). Rengasamy and Berchmans (2012) tried to feed MFCs with bad wine; several double chambered MFCs were built and operated, using different microbial inocula: Acetobacter aceti (A-MFC), Gluconobacter roseus (G-MFC), and also using a mixed culture (AG-MFC). The best experimental results using bad wine as a substrate (batch mode, 72 h) were obtained in the mixed culture fuel cell (AG-MFC), that generated a power density of 3.8 6 0.2 W/m3 with 45% CE and 41% COD removal. Overall performance for single culture MFCs was lower in terms of power density and CE, while COD removal was higher (up to 59% for A-MFC). Increasing the period of operation from 72 to 144 h, COD removal in the same MFC reached 87.5%. Cusick et al. (2010) operated four single chamber—cubic-shaped MFC reactors for WW treatment fed in batch mode. Anodes were built out of graphite fiber brushes, the air-cathode was made of carbon cloth coated of a Pt catalyst (0.5 mg/cm2). Batch cycle time was 6 days, obtaining a total COD removal of 65% 6 7%, CE of 18% 6 4%, and an energy recovery of 0.26 6 0.01 kWh/kg COD. Pepe Sciarria et al. (2015) tested white and red wine lees as a substrate for two single-chamber, air-cathode MFCs. Anodes were made of a core of two titanium wires with carbon fiber, while cathodes were made of platinum-coated carbon cloth (0.4 mg Pt/cm2, BASF). White wine lees were more performing in terms of power obtained (263 mW/m2), reduction of TCOD (90%) and BOD5 (95%) and CE (15%), if compared to red wine lees (111 mW/m2, 27%, 83%, and 9% respectively). Different substrate composition influenced the system’s overall performance and led to the development of different microbial consortia within the cells. The presence of polyphenols was shown to negatively influence MFCs performance. A list of MFCs performance using beverage industry (brewery and winery) wastewater as a substrate is reported in Table 5.5.

Table 5.5 List of beverage industry wastewater used as anodic substrate in microbial fuel cells (MFCs) and their respective performances. Substrate

MFC type

Feeding mode

Volume (NAC)

Anode material

Cathode material

Substrate concentration

COD removal (ηCOD)

Power density (max)a

CE

Ref.

Beer brewery processing wastewater

Air-cathode MFC

Batch

28 mL

Carbon cloth (30% wet proofed) with Pt catalyst (0.35 mg/ cm2)

2240 mg COD/L

87%

483 mW/m2

38%

Feng et al. (2008)

Beer brewery wastewater

One-chamber air-cathode MFC

Continuous

100 mL

Carbon cloth without wet proofing (7 cm2) Three parallel groups of carbon fibers

625 mg COD/L

43%

264 mW/m2

19.75%

Wen et al. (2009)

Beer brewery wastewater

One-chamber air-cathode MFC

Continuous

100 mL

Three parallel groups of carbon fibers

1501 mg COD/L

47.6

669 mW/m2

2.5%

Wen et al. (2010a)

Brewery wastewater

Tubular aircathode MFC

Batch

170 mL

Graphite felt/ granular graphite

2125 mg COD/L21

93%

96 mW/m2

28%

Zhuang et al. (2010)

Brewery wastewater

Serpentine type stack MFC (40 aircathode units) Stackable baffled MFC (90 L)

Continuous

250 mL (unit)

Graphite felt

Catalyst layer (88 wt.% ACP and 12% PTFE, 0.8 mg/cm2 Pt catalyst) 1 stainlesssteel net 1 wet-proof gas diffusion layers Catalyst layer (88 wt.% ACP and 12% PTFE, 0.8 mg/cm2 Pt catalyst) 1 stainlesssteel net 1 wet-proof gas diffusion layers GORE-TEX cloth with a mixture of Ni-based conductive paint and an ORR catalyst GORE-TEX cloth with a mixture of Ni-based conductive paint and an ORR catalyst

2120 mg COD/L

86.4%

97.2 mW/m2

7.6%

Zhuang et al. (2012)

Continuous

900 mL (unit)

Carbon brushes

800 mg COD/L

87.6%

181 6 21 mW/m2

19.1%

Dong et al. (2015)

Brewery wastewater

Rolling cathode made of activated carbon and PTFE

(Continued)

Table 5.5 (Continued) Substrate

MFC type

Feeding mode

Volume (NAC)

Anode material

Cathode material

Substrate concentration

COD removal (ηCOD)

Power density (max)a

CE

Ref.

Brewery wastewater (diluted)

Two 10 L tubular dualchamber MFCs SMFC

Batch (60 days) Continuous (325 days)

9.4 L

Carbon fiber cloth (1.51 m2)

Carbon fiber cloth (2.59 m2)

3197 6 979 mg COD/L

75.3% 94.6%

1.61 mW/m2

5.5% 13.9%

Lu et al. (2017)

Batch

225 mL

Graphite felt (24 cm2)

510 mg COD/L

b

890 6 18 mW/m2

57% 6 7%

Yu et al. (2015)

Batch

125 mL

Carbon felt (60 cm2)

Carbon cloth (30% wet proof) with Pt catalyst Graphite (60 cm2)

7.8 g COD/L

41%

3.82 W/m3

45%

Batch

28 mL

Carbon brush (0.22 m2)

10,100 mg COD/L

90%

263 mW/m2

15%

Semicontinuous

70 mL

Carbon felt

Carbon cloth (30% wet proofed) 1 Pt catalyst (0.4 mgPt/ cm2) Carbon felt

Rengasamy and Berchmans (2012) Pepe Sciarria et al. (2015)

6850 mg COD/L

17%

465 mW/m2

15%

65%

2

18%

Brewery wastewater (diluted) Reject wine

White and red wine lees

Winery wastewater Winery wastewater

Twochambered MFC Air-cathode MFC

Dual-chamber MFC Twochambered MFC

Batch

28 mL

Graphite fiber brush

COD, Chemical oxygen demand; SMFC, single-chamber MFC; ACP, activated carbon powder. a If not otherwise specified, power density is normalized to anodic chamber volume (W/m3) or anode electrode surface (mW/m2). b Indicates the data not available from the cited reference.

Carbon cloth 1 PTFE 1 Pt catalyst (0.5 mg/cm2)

2200 mg COD/L

278 mW/m 31.7 Wh/m3

Penteado et al. (2016) Cusick et al. (2010)

Agro-industrial wastewater treatment in microbial fuel cells

5.5

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Agro-industrial wastewaters and by-products

Agricultural activities may lead to the production of food, biofuels, and residues (wastes) in variable quantities from a large number of unit processes, from individual ones to the mixing of multiple effluent streams. Wastes and wastewater from production streams require treatment to avoid water pollution, keeping in mind that wastewaters vary strongly in composition and concentration of pollutants, not only between different regions and agro-industries, but also within individual production facilities. MFCs have been tested to treat some of these waste streams with promising characteristic.

5.5.1 Palm oil industry wastewater Palm oil is an edible oil produced mainly in tropical countries for food and energy production. Palm oil is a perennial crop in tropical regions; therefore, palm oil production, unlike olive oil, for example, is constant throughout the year. The two main products of this industry are crude palm oil (CPO) and the solid palm kernel (Garcia-Nunez et al., 2016). Most of the CPO is produced by wet process, which consists of sterilization, digestion, and oil-extraction steps. However, palm oil agroindustry produces also a significant amount of residual biomass of different types (Ng et al., 2012); amongst the many types of residuals, palm oil mill effluent (POME) results from the process of palm oil extraction in palm oil mills. No added chemicals can be detected in the POME because only water is used in the process. In general, 1 t of fresh fruit bunches (FFB) generates 0.56 m3 of POME with a content of 29.90 kg BOD/t FFB, 70.71 kg COD/t FFB, 12.8 kg TSS (total suspended solids)/t FFB and a content of oil and grease of 8.15 kg/ton FFB (Prasertsan et al., 2009). The use of residual biomass from the palm oil agro-industry has gained attention in recent years since it can be converted through the use of different technologies into value-added products (Chiew and Shimada, 2013; Garcia-Nunez et al., 2016). POME is characterized by complex substrates comprising amino acids, inorganic nutrients, and a mixture of carbohydrates ranging from hemicelluloses to simple sugars. More than 2.5 t of POME can be produced in the processing of 1.0 t CPO (Ahmad et al., 2003). POME is a liquid waste composed of about 94 wt.% of water and 6 wt.% solids (soluble or insoluble), which are essentially organic (Garcia-Nunez et al., 2016). Typically, COD and BOD are up to 50,000 and 25,000 mg/L, respectively. In addition, its high acidic content may cause environmental concern. In Table 5.6, chemical physical characterization of POME is reported. Conventionally, such residuals are handled successfully through anaerobic digestion processes in which B90% of the organic content is eliminated and simultaneously energy-rich methane gas is produced (Kang et al., 2017). Conventional methods required for initial POME hydrolysis are highly energy consuming, as high levels of suspended solids in the effluent are either floated or settled during treatment. The flotation in the form of scum on the surface of the anaerobic pond is

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Table 5.6 Chemical physical characterization of palm oil mill effluent. Parameter

U.M.

Concentration range

TSS COD Biological oxygen demand (BOD5) Fats and acids Ammonium (NH3-N) Total nitrogen (Ntot) Phosphate (P) pH

mg SST/L mg O2/L mg O2/L mg/L mg NH3-N/L mg N/L mg P2O5/L

5000 88,258 39,650 113,000 18,700 175,521 1963 80,701 4 80 180 1400 180 368 3.4 5.3

COD, Chemical oxygen demand; TSS, total suspended solids. Source: Adapted from Garcia-Nunez, J.A., Ramirez-Contreras, N.E., Rodriguez, D.T., Silva-Lora, E., Frear, C.S., Stockle, C., et al., 2016. Evolution of palm oil mills into bio-refineries: literature review on current and potential uses of residual biomass and effluents. Resour. Conserv. Recycl. 110, 99 114. https://doi.org/10.1016/j. resconrec.2016.03.022 and Madaki, Y.S., Seng, L., 2013. Palm oil mill effluent (Pome) from Malaysia palm oil mills: waste or resource. Int. J. Sci. Environ. 2(6), 1138 1155.

partially the result of enzymatic reactions (e.g., xylanase), which set the oil droplets free and then move together with the solids to the surface. The high suspended solids can also reach the surface due to the gas flow produced during anaerobic digestion. The sedimentation of solids reduces the capacity of the vessel, and the frequent sediment clean-up inevitably increases the cost of wastewater treatment. Treated wastewater is brownish, and thus not allowed to be discharged into the natural waterways, despite BOD values mostly within limits set by regulations. This gives no alternative other than reserving a very large area for the wastewater treatment system merely as reservoir (Fig. 5.3). Treatment by MFCs may, on the other hand, positively intervene on the overall energy balance due to the power produced in the bioelectrochemical process (Cheng et al., 2010; Lean˜o et al., 2012; Baranitharan et al., 2015). Bioelectrochemical recovery of electricity from POME by MFCs has thus been investigated in recent years. Double-chamber MFCs have been studied to treat raw POME inoculated with anaerobic sludge. The process showed a Pdmax (maximum power density) and volumetric power density of about 45 mW/m2 and 304 mW/m3, respectively, but low CE (0.8%) and COD removal efficiency (45%) were obtained (Baranitharan et al., 2013). In a follow-up study, however, Baranitharan et al. (2015) reported a more encouraging Pdmax of 107.35 mW/m2 with POME feed. In 2010 another group of researchers examined POME treatment using a two-stage MFC system integrated with immobilized biological aerated filters. It resulted in Pdmax of about 44.6 mW/m2 and more than 90% of total COD removal (Cheng et al., 2010). Kang et al. (2017) studied a dual-chamber MFC (V 5 150 mL) equipped with anode enhanced through the utilization of a conductive polymer (poly 3,4-ethylenedioxythiophene—PEDOT), using diluted POME (COD 350 mg/L) and acetate, at different percentages (25% 50% 75% 100%). Maximum power density achieved was 1.1 W/m2, CE of 24.3%, and ηCOD of 76.3%, with 50% POME solution as influent.

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Figure 5.3 Wastewater discharge of POME effluent in Malaysia. POME, Palm oil mill effluent.

Tee et al. (2018) tested a single-chamber MFC integrated with adsorption system (MFC-AHS) under various operating temperatures with POME as a substrate. The optimal operating temperature for such system was found to be at B35 C with power density, CE and maximum COD removal of 74 6 6 mW/m3, 10.65% 6 0.5%, and 93.57% 6 1.2%, respectively. MFC-AHS has demonstrated higher COD removals when compared to standalone MFC regardless of operating temperatures throughout the experiment. Tan et al. (2017) studied an integrated MFC-MBR system for POME treatment. The integration of MBR with MFC forms a bioelectrochemical membrane reactor, which takes advantage of both MBR and MFC properties, enhancing effluent quality while also achieving better energy recovery. In this study an MFC was incorporated into the anaerobic membrane bioreactor (AnMBR) system (MFC-AnMBR), acting as a pretreatment unit prior to the AnMBR. Temperature variations and their impact on the degree of membrane fouling was also observed, with the AnMBR operated at mesophilic temperature, 45 C, being the best performer in COD removal efficiency (97.66%). By combining MFCs with AnMBRs, overall observed COD removal efficiency was higher compared to the application of AnMBRs alone. Introduction of MFCs as a pretreatment to AnMBR helped reduce the presence of fine foulants and filamentous bacteria and thus also achieve better filtration performance. A summary of different experiments using POME wastewater as a substrate is reported in Table 5.7.

5.5.2 Agricultural products processing wastewater Agricultural products processing industries are amongst the most water-consuming; they include wheat starch, potato and sweet potato starch, mustard tuber, olive mill

Table 5.7 List of palm oil mill effluent wastewater used as substrate in microbial fuel cells (MFCs) and their respective performances. Substrate

MFC type

Feeding mode

Volume (NAC)

Anode material

Cathode material

Substrate concentration COD removal (ηCOD)

Power density (max)a

Palm oil mill effluent

Dual-chambered MFC

Batch

450 mL

PACF 59 cm2

PACF 59 cm2

2680 mg COD/L

32%

107.35 mW/m2 74%

Palm oil mill effluent

MFC (pretreatment) 1 AnMBR Cylindrical MFC

Continuous

b

b

b

19,635 6 923 mg COD/L

77.1% 6 1.1%

b

b

Continuous

b

PbO2 electrode 8000 10,500 mg COD/L 90% 1 copper wire

44.5 mW/m2

b

Dual-chambered MFC

Batch

150 mL

Granular graphite 1 graphite rod/ carbon fiber felt 1 copper wire Graphite felt (32 cm2)

Carbon cloth

b

76.3%

1.1 W/m2

24.3% Kang et al. (2017)

Dual-chambered MFC

Batch

1.8 L

Aluminum rod

Copper rod

38,400 mg COD/L

54%

18.92 mW/m2

b

Palm oil mill effluent

Palm oil mill effluent (diluted) 1 acetate Palm oil mill effluent

AnMBR, Anaerobic membrane bioreactor; COD, Chemical oxygen demand; PACF, polyacrylonitrile carbon felt. a If not otherwise specified, power density is normalized to anodic chamber volume (W/m3) or anode electrode surface (mW/m2). b Indicates the data not available from the cited reference.

CE

Ref.

Baranitharan et al. (2015) Tan et al. (2017) Cheng et al. (2010)

Lean˜o et al. (2012)

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111

wastewater (OMW), cereal processing wastewater, and molasses-based industry for sugar production. Starch products manufacturing industry utilizes large amounts of water, resulting consequently in large volumes of starch processing wastewater (SPW) characterized by high COD, ranging from 16,870 to 22,800 mg COD/L. SPW contains a relatively high content of carbohydrates (2300 3500 mg/L), sugars (0.65% 1.18%), protein (0.12% 0.15%), and starch (1500 2600 mg/L), representing both an important energy-rich resource, potentially convertible to a wide variety of useful products and a possible threat to the environment due to its composition (Jin et al., 2002). Cassava sour starch is the product of traditional rural low-technology agro-industry (Colin et al., 2007). Cassava processing generates two liquid waste streams: the first, resulting from the cassava roots washing and peeling in a rotary drum, generally contains large amounts of inert materials with low COD; the latter, resulting from drainage of the starch sedimentation tank, with higher COD and BOD concentrations. Wastewater characteristics are highly dependent on the efficiency of the equipment used in the factory. In a typical family-run sour starch plant, 1 5 t of cassava roots are processed daily and about 12 m3 of water are consumed per ton of fresh cassava roots (Colin et al., 2007). In the starch-production process, a ton of fresh cassava root can generate about 0.2 t of starch, 0.4 0.9 t of residue, and 5 7 L of wastewater. Cassava mill wastewater is a carbohydrate-rich starch waste having high COD, BOD, TS, and low ammonium nitrogen concentrations (Kaewkannetra et al., 2011), with high cyanide content (average 3.5 mg/L) and acid (pH 4.5 5.5) (Colin et al., 2007). Starch and cassava processing wastewater chemical physical characterization is reported in Table 5.8. Mustard tuber wastewater (MTWW) processing is a high-strength and highsalinity effluent often generated in large volumes. Guo et al. (2013) reported using a dual-chamber MFCs on MTWW, obtaining a Pdmax of 246 mW/m2, 67% and 85% CE, and COD removal, respectively. Molasses obtained from sugarcane mills are widely used in fermentation industries, and they represent one of the most important raw materials for ethanol production due to the low cost and wide availability. Molasses-based wastewater characterization is reported in Table 5.9. Table 5.8 Chemical physical characterization of starch processing wastewater Parameter

U.M.

Starch

Cassava

TSS COD Biological oxygen demand (BOD5) Ammonium (NH3-N) Total Nitrogen (Ntot) Phosphate (P) pH

mg SST/L mg O2/L mg O2/L mg NH3-N/L mg N/L mg P2O5/L

2600 4400 16,870 22,800 9400 13,200 440 620 500 28,000 92 190 4.4 5.4

700 3500 4200 17,000 1100 3900 35 39 80 370 20 35 3.6 6.5

COD, Chemical oxygen demand; TSS, total suspended solids. Source: Based on Colin et al., 2007; Jin et al., 2002; Kaewkannetra et al., 2011.

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Table 5.9 Chemical physical characterization of molasses-based distillery wastewater Parameter

U.M.

Cane

Beet

TSS COD Biological oxygen demand (BOD5) Ammonium (NH3-N) Total nitrogen (Ntot) Phosphate (P) pH

mg SST/L mg O2/L mg O2/L mg NH3-N/L mg N/L mg P2O5/L

2600 4400 16,870 22,800 9400 13,200 440 620 500 28,000 92 190 4.4 5.4

700 3500 4200 17,000 1100 3900 35 39 80 370 20 35 3.6 6.5

COD, Chemical oxygen demand; TSS, total suspended solids. Source: Based on Satyawali and Balakrishnan, 2008.

The use of molasses as raw materials for producing fermentation products (i.e., alcohol and amino acids) produces a large amount of high-strength wastewater. Several chemical and biological methods for molasses treatment have been tested, and MFC is amongst them. However, as reported by Zhang et al. (2009a), Mohanakrishna et al. (2010), and Sevda et al. (2013), the complex nature of the substrate makes the simultaneous accomplishment of wastewater treatment and energy recovery difficult. Crude starch extract from potatoes was used as a substrate for inoculation (instead of glucose) in a two-chambered, mediator-less MFC by Herrero-Hernandez et al. (2013), achieving a CE and ηCOD of 18.5% and 61%, respectively. SPW was used by Lu et al. (2009) as anodic substrate in an air-cathode MFC operated over four cycles (140 days). The MFC achieved COD and ammonia nitrogen removal of 98% and 90.6%, respectively, and a maximum CE of 8%. Amongst other agro-industrial effluents, OMW, generated during the production of olive oil, is one of the most hardly polluted wastewater, reaching COD values up to 100,000 mg/L. Olive oil is one of the main agricultural products in the Mediterranean area, contributing for 95% to the annual worldwide olive oil production (over 30 million m3/year) (Pepe´ Sciarria et al., 2013). Italy is one of the main olive oil producers in the world (Source: EUROSTAT, 2004 07). About 2 3 106 m3 OMW are produced annually in Italy. The volume of these effluents has increased markedly in the last decades. The treatment of olive mill effluents includes wastewater generated in the washing of the olives, olive washing wastewater (OWW), as well as OMW, wastewater from olive oil washing, as well as from other activities in the facility, including cleaning and sanitation. OWW is characterized by a high concentration of suspended solids (mainly, peel, pulp, ground, branches, and leaves debris) dragged during the olive fruit washing process but also by a low concentration of dissolved organic matter—depending on the water flow exchange rate in the washing machines during the fruit cleaning procedure—usually lower than the limits set by national standards for the discharge of the effluent on superficial land.

Agro-industrial wastewater treatment in microbial fuel cells

113

OMW is characterized by strong odor nuisance, acid pH, intensive black color, and toxic properties [due to the presence of polyphenols, in a 200 8000 mg/L range (Azbar et al., 2004)], exhibiting considerable electroconductivity values (Ochando-Pulido and Martinez-Ferez, 2017). In addition, OMW contains free fatty acids and carbohydrates. Uncontrolled disposal of these effluents represents an environmental hazard, and an appropriate treatment is required before discharge (Danellakis et al., 2011). Due to the effluent characteristics (high COD load, presence of refractory compounds, fats, and lipids) untreated direct discharge of these wastewaters to the municipal sewage treatment plants is forbidden (Ochando-Pulido and Martinez-Ferez, 2017) since some substances (especially polyphenols) can potentially have inhibitory effects on bacteria involved in OMW biological treatment (Ntaikou et al., 2009). OMW disposal generally requires physicochemical treatment, such as coupled evaporation and combustion, chemical coagulation and sorption, and oxidation (Kestio˘gǧlu et al., 2005), although biological treatments, such as anaerobic digestion, have also been tested (Bertin et al., 2010). A Fenton-like oxidational process pretreatment has been proposed to reduce the toxicity of olive oil wastewater before anaerobic digestion (Kiril Mert et al., 2010), while chemical methods are neither environment friendly, nor cheap. Commonly, dilution of wastewater is necessary to eliminate microbial inhibition in biological processes (Azbar et al., 2004). Bermek et al. (2014) fed a small single-chamber air-cathode MFC (12 mL inner volume) with diluted OMW (1:10) as sole carbon source and obtained COD and total phenolics removal of 65% and 49%, respectively, with a maximum voltage generation of 381 mV. Pepe´ Sciarria et al. (2013) operated a batch-fed single-chamber air-cathode MFC containing graphite fiber brush anode and a cathode equipped with Pt-coated (0.5 mgPt/cm2) carbon cloth for OMW treatment. Despite the fact that OMW is not considered a good substrate for microbiological process due to its characteristics, the mixture of domestic wastewater and OMW (14:1) achieved a good power generation (124.6 mW/m2), COD removal (60%) and CE (29%), and highlighted that OMW mixed with domestic wastewater could be suitably treated in this unconventional process. Table 5.10 summarizes a list of agricultural products processing wastewater used as substrate in MFCs and their respective performances.

5.5.3 Agricultural residues The abundance and renewability of cellulosic and lignocellulosic materials from agricultural residues makes them a cheap renewable energy and carbon resources. However, lignocellulosic biomass cannot be directly utilized by microorganisms in MFCs for electricity generation, while fermentable sugars can be derived from these materials by acid or enzymatic treatments (ElMekawy et al., 2015). Wheat straw is an agricultural residue. The organic carbon content in wheat straw is constituted of about 34% 40% cellulose, 21% 26% hemicellulose and 11% 23% lignin content, usually it is hydrolyzed to obtain a carbohydraterich liquid substrate, called wheat straw hydrolysate (Khan and Mubeen, 2012).

Table 5.10 List of agricultural products processing wastewater used as substrate in microbial fuel cells (MFCs) and their respective performances. Substrate

MFC type

Feeding mode

Beet-sugar wastewater

Upflow anaerobic sludge blanket reactor MFC (air-cathode MFC) ABSMFC MFC (4 units)

Anode material

Cathode material

Substrate concentration

COD removal (ηCOD)

Power density (max) a

CE

Ref.

Continuous 650 mL

Granular graphite (100 g) 1 graphite rod

Carbon paper 1 PTFE 1 Pt catalyst (0.5 mg/ cm2)

127,500 mg COD/L (diluted 20 5 0 times)

53.2%

1410 mW/m2

1%

Zhang et al. (2009a)

Continuous 690 mL

Carbon fiber felt (56 cm2)

127,500 mgCOD/L (diluted 40 20 10 times)

70%

115.5 mW/m2

b

Zhong et al. (2011)

Two-chambered MFC Dual-chambered

Continuous 15 L

16,000 mg COD/L

72%

1771 mW/m2

20%

Batch

310 mL

Graphite plate (0.16 m2) Carbon paper

595 mg COD/L

95 %

81 mW/m2

Distillery wastewater (molasses based) Molasses wastewater mixed with sewage

Single-chamber open aircathode MFC

Batch

500 mL

Carbon cloth (wet proofed) 1 PTFE 1 Pt catalyst (0.5 mg/cm2) Graphite plates (0.16 m2) Carbon paper 1 Pt catalyst (0.5 mg/ cm2) Graphite plate (70 cm2)

Cassava mill wastewater Cereal processing wastewater

15.2 kg COD/(m3 day)

72.8%

124.35 mW/m2

b

Mohanakrishna et al. (2010)

Single-chambered MFC

Continuous 25 mL

9958 mg COD/L

59%

382 mW/m2

b

Sevda et al. (2013)

Mustard tuber wastewater Olive mill wastewater (diluted 1:10)

Dual-chambered MFC Single-chamber air-cathode MFC

Batch

550 mg COD/L

57.1%

246 mW/m2

65%

b

Beet-sugar wastewater

Batch

Volume (NAC)

150 mL 16 mL

Graphite plate (70 cm2)

graphite fiber brush Pressed AC 1 PTFE 1 Pt (10 cm2) 1 catalyst 1 stainlessstainless-steel steel mesh current mesh current collector collector Carbon cloth Carbon cloth (40.5 cm2) (40.5 cm2) Carbon cloth (7 cm2) Carbon cloth (7 cm2)

20.9 g COD/L

Kaewkannetra et al. (2011) 40.5% Oh and Logan (2005)

67.7% Guo et al. (2013) Bermek et al. (2014)

b

Olive mill wastewater 1 Domesti wastewater (1:14) Starch extract (potatoes) Starch processing wastewater

Carbon cloth (30 wt.% wet proofed) 1 PTFE 1 Pt catalyst (0.5 mg/cm2)

4000 6 410 mg COD/L 60%

124.6 mW/m2

29%

1000 mL Titanium mesh (6.5 cm2)

Titanium mesh (6.5 cm2)

552 2677 mg COD/L

61%

502 mW/m2

b

Carbon paper with 1.12 mg/cm2 Pt catalyst (17 cm2)

4852 mg COD/L

98.0%

239.4 mW/m2

18.5% HerreroHernandez et al. (2013) 8% Lu et al. (2009)

Single-chamber air-cathode MFC

Batch

28 mL

Mediator-less, two-chambered MFC Air-cathode single-chamber MFC

Batch

Batch

Brush anodes (0.22 m2)

Carbon paper (25 cm2)

COD, Chemical oxygen demand. a If not otherwise specified, power density is normalized to anodic chamber volume (W/m3) or anode electrode surface (mW/m2). b Indicates the data not available from the cited reference.

Pepe´ Sciarria et al. (2013)

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Integrated Microbial Fuel Cells for Wastewater Treatment

This substrate has been used for electricity generation with MFCs in a few studies. Hydrolysate obtained after liquefaction of wheat straw has been used in an H-type, two-chambered MFC after dilution with wastewater. Maximum power density reached was 123 mW/m2 when initial COD of hydrolysate was 1 g/L (Zhang et al., 2009a,b). Though power density reported was low, its use showed suitability as a substrate for MFC. Corn stover is an agricultural by-product mainly composed of cellulose and hemicellulose (70%), which needs to be converted into sugars by cellulosic enzyme treatment or steam explosion (ElMekawy et al., 2015). Raw corn stover was fed experimentally as a substrate for electricity generation in a singlechamber MFC, proving that the treatment was effective (Wang et al., 2009), though the power output was much lower than a control-MFC run with glucose as substrate. Zuo et al. (2006) used corn stover hydrolysate of 1 g COD/L from neutral and acid steam exploded hydrolysis in single-chamber air-cathode MFC, obtaining power densities of 371 6 13 and 367 6 13 mW/m2, respectively, and 93% 95% BOD removal. Power densities from corn stover hydrolysate can be increased by improving cathode architecture and also by increasing solution conductivity. Increasing the conductivity of electrolytes to 20 mS/cm, Zuo et al. (2006) obtained 861 6 37 mW/m2 with acid hydrolysate. Breaking down cellulose directly into sugar using specialized microorganism or consortium is useful to produce the desired product in a single step process. Wang et al (2009) produced electricity directly from powdered raw corn stover with a bottle-type air-cathode MFC, using an acclimated mixed culture for cellulose breakdown together with exoelectrogenic bacteria. A maximum power density of 331 mW/m2 was produced from raw corn stover, with a COD removal efficiency of 48%. However, the energy recovery from raw corn stover was very low (3.6%). A multistep consortium could be useful to obtain corn stover saccharification and electricity generation. Rice straw is one of the cheapest and most abundant agricultural wastes. Rice straw is mainly composed of cellulose, hemicellulose, and lignin. Electricity generation from rice plants can be achieved in MFCs using rice hydrolysate as a substrate, as well as raw rice straw directly placed in MFCs with cellulose-degrading anaerobes. Rice milling industry wastewater maximum sustainable volumetric Pd and COD removal of 2.3 W/m3 and 96.5%, respectively, were reported using earthen pot MFC, with an additional 84% lignin and 81% phenol removal (Behera et al., 2010). The maximum power density of 137 mW/m2 was obtained from rice straw hydrolysate of 400 mg COD/L and increasing the solution conductivity to 17 mS/cm, the power density reached 293 mW/m2 (Wang et al., 2014a,b). COD removal efficiency ranged from 49.4% to 72%, whereas CE varied from 8.5% to 17.9%. Direct use of powdered rice straw as anodic substrate in a two-chambered MFC without pretreatment (1 g COD/L) was also tested with a mixed culture of cellulose-degrading bacteria (Hassan et al., 2014), obtaining a CE up to 54.3%. Gurung and Oh (2015) used rice straw without pretreatment and inoculated with a mixed culture of cellulose-degrading bacteria in MFC. At an initial concentration of 1 g/L, Pdmax of 190 mW/m2 was reported.

Agro-industrial wastewater treatment in microbial fuel cells

117

So far, MFCs have proven to be effective in treatment of waste materials from rice cultivation and harvesting for bioelectricity production. Table 5.11 summarizes a list of agricultural residuals used as substrate in MFCs and their respective performances.

5.6

Livestock industry wastewater

The livestock industry includes slaughterhouses and manure processing. Meatprocessing activities use plenty of water for hygienic reasons and thus produce a large amount of wastewater. The existence of large amounts of suspended solids in these wastewaters generates odors, a major environmental problem associated with this type of waste. Slaughterhouses produce high-strength wastewater, mainly constituted of biodegradable organic carbon, fats, and proteins, characterized by the presence of high concentrations of animal blood, skinning residuals, and washwater from cleaning of animal carcasses and ambients. Slaughterhouse wastewater have high organic content and suspended solids, and high concentration of nutrients (Sunder and Satyanarayan, 2013); their quality depends on the efficiency of blood retention during animal bleeding (to reduce BOD), water usage (water economy increase pollutant concentration, although total mass will remain constant), type of animal slaughtered (BOD is higher in cattle wastewater than swines’), meatprocessing activities (plants that only slaughter animals produce stronger wastewater than those involved in rendering or other meat-processing activities) (Masse´ and Masse, 2000). Chemical physical characterization of slaughterhouse wastewater is reported in Table 5.12. Wastewater discharge without proper treatment from slaughterhouses may deeply contaminate water bodies due to its high organic content. Animal carcass wastewater is conventionally disposed of through anaerobic treatment and alkaline hydrolysis to accelerate the degradation process and reduce its environmental impact. The sterile solution produced, that is, animal carcass wastewater, is a coffee-colored alkaline liquid characterized by high amounts of BOD (70 g/L), COD (105 g/L), ammonia (1 g/L), organic nitrogen (8 g/L), and TP (0.4 g/L) (Das, 2008). The possibility of using both slaughterhouse and animal carcass wastewater in MFCs has been reported. For instance, slaughterhouse wastewater in dualchambered MFC resulted in Pdmax of 578 mW/m2 (Katuri et al., 2012). Bioelectricity generation with animal carcass wastewater was evaluated using upflow tubular air-cathode MFCs. It resulted in Pdmax of 2.19 W/m3 with 50.66% COD removal, but with very low CE (0.25%) (Li et al., 2013). Animal manure wastewater has been thought to be the primary element generating wastes from agricultural activity. In the livestock industry, swine wastewater and cattle manure wastewater are the most widely used. Animal manure wastewater-treatment technology is important to achieve sustainable animal production. In methane gas generation during anaerobic treatment, generated ammonia and odor (due to volatile organic acids in the wastewater) are a cause of concern

Table 5.11 List of agricultural residues used as anodic substrate in microbial fuel cells (MFCs) and their respective performances. Substrate

MFC type

Feeding mode

Volume (NAC) (mL)

Anode material

Cathode material

Substrate concentration

COD removal (ηCOD)

Power density (max) a

CE

Ref.

Powdered rice straw

Twochambered MFC H-type MFC

Batch

160

Carbon paper (10 cm2)

Carbon paper (10 cm2)

1000 mg COD/L

b

190 mW/m2

37%

Gurung and Oh (2015)

Batch

200

Carbon paper (10 cm2)

Carbon paper (10 cm2)

1000 mg COD/L

b

145 mW/m2

54.3% 45.3 %

Batch Singlechamber aircathode MFC Bottle-type Batch aircathode MFC Earthen pot Batch MFC

28

Carbon paper (7.1 cm2)

Carbon cloth 1 Pt catalyst (0.5 mg/cm2) (7.1 cm2)

1 g COD/L

60% 70%

861 6 37 mW/m2

20% 30%

Hassan et al. (2014) Zuo et al. (2006)

250

Carbon paper

b

3.6%

Wang et al. (2009)

2250 mg COD/L

2.3 W/m3

21%

Batch Aircathode singlechamber MFC Bottle-type Batch aircathode MFC Batch H-type dualchamber MFC

220

Stainless-steel mesh (190 cm2) Carbon brush

42% 6 8% (cellulose) 17% 6 7% (hemicellulose) 96.5%

331 mW/m2

400

Carbon cloth 1 Pt catalyst (0.35 mg/cm2) (4.9 cm2) Graphite plate (231 cm2) Gas diffusion electrode 1 Pt catalyst (0.5 mg/ cm2)

400 mg COD/L

49% 72%

137.6 6 15.5 mW/ 8.5% 17% m2

Behera et al. (2010) Wang et al. (2014b)

250

Carbon paper

Carbon cloth 1 Pt catalyst (0.35 mg/ cm2) (4.9 cm2)

b

60% 6 4% (cellulose) 15% 6 4% (hemicellulose)

406 mW/m2

1.6%

Wang et al. (2009)

300

Carbon paper (42 cm2)

Carbon paper (42 cm2)

250 2000 mg COD/L

b

123 mW/m2

15.5% 37.1%

Zhang et al. (2009b)

Powdered rice straw Steam exploded corn stover

Raw corn stover

Rice milling

Rice straw hydrolysate

Steam exploded corn stover residue Wheat straw hydrolysate

COD, Chemical oxygen demand. a If not otherwise specified, power density is normalized to anodic chamber volume (W/m3) or anode electrode surface (mW/m2). b Indicates the data not available from the cited reference.

Agro-industrial wastewater treatment in microbial fuel cells

119

Table 5.12 Chemical physical characterization of slaughterhouse wastewater Parameter

U.M.

Concentration range

TSS COD Biological oxygen demand (BOD5) Ammonium (NH3-N) TKN pH

mg SST/L mg O2/L mg O2/L mg NH3-N/L mg N/L

10,120 14,225 6185 6840 3000 3500 650 735 1050 1200 8.0 8.5

COD, Chemical oxygen demand; TKN, total Kjeldahl nitrogen; TSS, total suspended solids. Source: Based on Alayu and Yirgu, 2018.

(Chen et al., 2018; Pandey et al., 2016). Animal manure is normally considered high-strength wastewater, rich in nitrates and phosphates, and must be treated to meet discharge regulations to avoid water contamination and odor problems. Use of alkaline-thermal pretreated swine wastewater reduce COD, ammonium, and turbidity, and it was reported being a suitable preprocessing in enhancing MFC performance (Pandey et al., 2016); such a process can be considered a possibility to achieve simultaneously enhanced resource recovery, energy recovery, and swine wastewater treatment (Guo and Ma, 2015). Animal farms may also produce large amounts of cattle manure and manure wash-water during its operations. Cattle manure is a source of carbon and nitrogen, and it is characterized by the presence of some easily degradable compounds as well as cellulose, hemicellulose, and lignin, which are difficult to hydrolyze (Inoue et al., 2013). This complexity of substrates helps the proliferation of complex microbial communities. Therefore cattle manure has been used as an inoculum for MFCs, as well. Since exoelectrogenic bacteria can generally utilize a limited range of substrates, more diverse microbial communities are required to oxidize more complex organic matters (Kiely et al., 2011). The combined action of multiple species of microbes in a consortium has been reported to be necessary for producing greater current densities in MFCs. Cattle manure sludge was treated in an air-cathode MFC, with and without any mediators in which power density was increased by nearly 200% by using methylene blue as a mediator (Scott and Murano, 2007). Using a cassette-electrode MFC configuration in fed-batch mode, a maximum of 16.3 W/m3 (765 mW/m2) was reported with suspended cattle manure as substrate, achieving the removal of 41.9% of the total COD removal in first 10 days of MFC operation (Inoue et al., 2013). MFC fed with particulate cattle manure and manure wash-water generated power density of 67 215 mW/m2 in batch mode, which was comparable with results obtained from the liquid feedstock. A power yield of 15.1 W/m3 was reported when cattle manure was used as a suspended solid substrate in MFC (Zhang et al., 2012).

120

Integrated Microbial Fuel Cells for Wastewater Treatment

Cow manure (80% moisture) was used as organic carbon source in a compost MFC, producing a power density of 349 6 39 mW/m2 with 0.1 mg/cm2 Pt catalyst at the cathode (Wang et al., 2014a,b). Electricity production from cattle manure was limited to B10 kJ/kg wet manure, which can be converted into 50 kJ/kg dry manure (Zheng and Nirmalakhandan, 2010). The integration of MFC in manure and manure wash-water treatments was beneficial in terms of effective degradation of substrate, but the energy generated was limited in comparison to other bioenergy technologies. Preliminary tests using swine wastewater in two-chambered aqueous cathode MFC fed-batch mode showed a Pd of 45 mW/m2 (Min et al., 2005). More extensive tests of the same group with single-chambered air-cathode MFC produced a Pdmax of 261 mW/m2 (Jung et al., 2008) and 382 mW/m2 (Cheng et al., 2014). A summary of different experiences of livestock industry process wastewater used as an anodic substrate in MFCs and their respective performances is reported in Table 5.13.

5.7

Challenges in using microbial fuel cells

Evolution of MFC technology on many fronts (reactor designs, materials, substrates, biocatalysts amongst all) has brought it much closer to realize its full potential and application for bioenergy production and simultaneous wastewater treatment. The use of wastewaters as an electron donor is desirable due to the growing demand for ecological wastewater treatment with minimum carbon output and maximum energy recovery. Although power densities form MFC have risen by several orders of magnitude in the last decades, and some companies even launched MFC-based wastewater treatment systems, many challenges still remain. High operating costs and low power output should be considered before the commercial success of the MFC technology, both as an alone process or combined with other technologies. Besides the limitations and problems of current outputs, the selection of appropriate substrates in terms of molecular complexity and resistance is challenging. Wastewater concentration and composition are among the most important factors that affect MFCs performance. Low conductivity or pH far from the optimal range might slow down microbial activity; therefore, pH and its possible changes during operation must be considered in the substrate selection phase. In this chapter, many types of wastewater were analyzed as MFC feedstocks in terms of bioelectricity production and organic matter removal efficiency. The use of the cheap, available substrates, such as wastewater, as electron donors, is desirable due to the growing demand for ecological wastewater treatment with a minimum carbon footprint (Pandey et al., 2016). Among these, the distillery industry, along with agro-processing industry wastewaters, showed better efficiency because of the presence of methanogenic inhibitors and electron transferring mediators such as lignin already present in the substrate. Dairy and livestock industry wastewaters

Table 5.13 List of livestock industry process wastewater used as anodic substrate in microbial fuel cells (MFCs) and their respective performances. Substrate

MFC type

Feeding mode

Animal carcass wastewater

Upflow tubular air-cathode MFC

Cattle manure (suspended solid substrate) Cattle manure slurry

Cylindrical MFC

Air-cathode cassetteelectrode MFC Manure wash water Air-cathode singlechamber MFC Slaughterhouse Dualwastewater chambered MFC Swine wastewater Singlechambered MFC

Volume (NAC)

Anode material

Cathode material

Substrate concentration

COD removal (ηCOD)

Continuous 750 mL

Graphite felt 1 300 g GAC

11,180 mg COD/L 50.7%

Batch

617 mL

Graphite fiber brushes

Batch

550 mL

Carbon felt (94.5 cm2)

Carbon fiber cloth 1 PTFE 1 catalyst layer (MnO2) Granular graphite 1 graphite fiber brushes Carbon fiber cloth 1 PTFE (58.5 cm2)

Batch

b

Carbon brush

Batch

125 mL

Carbon cloth (20 cm2)

Batch

250 mL

Carbon paper

Power density (max) a

CE

Ref.

2.19 W/m3

0.25%

Li et al. (2013)

15.1 W/m3

39.8% 6 1.6%

Zhang et al. (2012)

b

39.6% 6 4.1%

b

41.9% 765 mW/m2 28.8% 56.7%

Inoue et al. (2013)

Air-cathode

b

b

215 mW/m2

Zheng and Nirmalakhandan (2010)

platinized titanium mesh cathodes (20 cm2) Carbon paper 1 Pt catalyst (0.35 mg/cm2)

4850 mg COD/L

93% 6 1%

578 mW/m2 64% 6 2%

Katuri et al. (2012)

8320 mg COD/L

b

182 mW/m2

Min et al. (2005)

COD, Chemical oxygen demand; GAC, granular acivated carbon. a If not otherwise specified, power density is normalized to anodic chamber volume (W/m3) or anode electrode surface (mW/m2). b Indicates the data not available from the cited reference.

b

b

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Integrated Microbial Fuel Cells for Wastewater Treatment

showed good performance, but they are limited by the presence of other electron acceptors and of nonexoelectrogenic microorganisms due to the complexity of the substrate, such as fermenters and methanogens. Animal processing industry wastewaters are efficient, especially for the presence of blood and organic compounds. Operational conditions such as flow mode, temperature, pH, OLR, hydraulic retention time, microbial selection, and internal resistance also play an important role in MFCs performance. Low CE has been a general issue in MFC systems fed with real wastewaters due to several factors: diversion of electrodes into nonexoelectrogenic biomass growth; substrate consumption via other competing metabolic processes, such as fermentation and methanogenesis, or presence of toxic/inhibitors for EABs; large number of electrons locked in substrates by other electron acceptors; and low electron transfer efficiency. Another challenge for MFC scaling-up (from laboratory scale to full scale) is building up economic and simple setups for wastewater treatment that could be easily maintained and produce high-power levels. The capital cost of MFCs, on average, is 30 times higher than that of a traditional activated sludge treatment system for domestic wastewater, because of their configuration and treatment capability (He et al., 2017). The high-capital cost in MFCs is due to the use of expensive electrode materials, catalysts, and membrane materials (Oliveira et al., 2013). New materials for MFCs are constantly explored and developed to improve their economic feasibility and performance. Scaling-up MFC technology from milliliters to liters generally led to higher potential losses at the surface of the electrodes, reducing the achievable current density (ElMekawy et al., 2015) considerably. The maximum energy recovered from BESs reported so far is quite low for their applicability: power density benchmarks of 1 kW/m3 were achieved so far only in small-scale systems (Arends and Verstraete, 2012). MFCs could be scaled-up using two different techniques: connection of single small cells together or amplification of a single cell volume (Cheng and Logan, 2011; Dewan et al., 2008). Increasing anode size may be beneficial. However, the cathode surface area has a critical role in scaling-up MFCs. Their performance is in fact influenced negatively by the decrease in the ratio between surface area and volume, with consequent power density drops, so the scaling-up of a single cell is very difficult to apply (Cheng and Logan, 2011; Dewan et al., 2008; ElMekawy et al., 2015). A scaled-up stack of MFC setup may be an effective alternative to overcome problems of electrode spacing, orientation, and surface area to obtain the best power outputs (Dewan et al., 2008; Ieropoulos et al., 2010). The application of stacked MFCs in parallel or in series would be essential to significantly increase bioelectricity generation. Specifically, the stacked MFCs in parallel can increase current, power density, and CE more than in series. On the other hand, stacked MFCs in direct series can increase voltage. Nevertheless, an unbalanced substrate distribution can lead to voltage reversal and ionic short circuits (Cecconet et al., 2018a,b). These are still huge barriers in the way of practical application that may result in a low electricity generation or even failure of the overall system. Electricity generation in a stacked MFC can

Agro-industrial wastewater treatment in microbial fuel cells

123

be greatly impacted by the electrode connection methods (in series or in parallel), hydraulic flow modes (in series or in parallel), as well as operating conditions.

5.8

Conclusion

This chapter summarized MFCs application results on substrates from different sources and compositions. Almost all these applications were limited to the lab or pilot scale, as there are no reports yet, about the application of this technology to real-scale water resource recovery facilities. This is due to a still limited technical understanding of the technology, especially concerning design parameter’s and material’s effects on full-scale system development. Initial costs and the limited levels of energetic recovery achieved so far, compared to theoretical expectations, seem to have deflected industrial interest away from this technology, even though scientific interest in it is still high, as demonstrated by intensive publication activities in this area. It should, however, be considered that MFC technology already offers, even at this development stage, several important advantages: low input energy costs (due to lack of aeration) and extremely low levels of residual biosolids production. So far, expensive solutions have been imagined maximizing energy recovery from these processes, however, their final balance could just as well turn out positive by forfeiting significant energy harvesting in favor of more economic constructive solutions, with sufficient pollutants removal capacity and low live operating costs, and the lack of high costs for the disposal of excess biosolids. The most interesting perspective so far for MFCs is the integration with existing treatment technology as a combined treatment to enhance the chances of their integration into existing wastewater treatment plants and to advance the prospects of MFCs for practical application. Substrates used in MFCs have grown in complexity if compared with the first experimental designs (where the only simple, synthetic substrate was used to evaluate the electricity production). The use of high-strength substrates (in terms of OLR) such as agro-industrial wastewater led to the development of a more complex and diverse electrochemically active biomass in the anodic chamber, improving the overall electric output of the system. However, electric outputs are still quite far from being appealing for large-scale system applications, especially in comparison with the cost of materials required to build MFC setups. The high capital cost required for building up this technology (already high at laboratory scale) is not attractive, despite the production of clean energy from waste materials.

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Pharmaceutical wastewater treatment in microbial fuel cell

6

Somdipta Bagchi and Manaswini Behera School of Infrastructure, IIT Bhubaneswar, Bhubaneswar, India

Chapter Outline 6.1 Introduction 135 6.2 Application to pharmaceutical wastewater treatment 138 6.3 Integration of microbial fuel cell with other wastewater-treatment processes 146 6.4 Large-scale microbial fuel cell: potentials and challenges 147 References 148

6.1

Introduction

With increasing population, urbanization, and industrialization the quantity of wastewater generation and need for renewable energy sources are increasing day by day. Thus to address both the issues, integrated waste-management technologies have to be designed, which can not only treat waste but also can generate energy from the waste. Bioelectrochemical systems (BESs) are a type of integrated wastemanagement technologies, which convert chemical energy stored in the bonds of organic waste into electrical energy, using microorganisms as catalyst. Microorganisms grow and degrade waste according to the prevailing microenvironment. In anaerobic environment, due to the absence of oxygen the terminal electron acceptor, electrons and protons formed as a result of degradation of organic, are accepted by the electrodes. Electrons, thus accepted by electrodes, then flow through wire generating current. BESs are usually classified into microbial fuel cell (MFC) and microbial electrolysis cell (MEC). A conventional dual chamber MFC consists of anode and cathode chamber separated by a semipermeable membrane. The waste is degraded anaerobically in anode chamber to generate electrons (e2), protons (H1), and CO2. These e2s and H1 flow through external wire and membrane to cathode, combining with oxygen to form water. In MEC an input of energy in cathode chamber facilitates the formation of H2 (Figs. 6.1 and 6.2). MFC was first invented by Potter (1911), who generated electricity from a laboratory culture of Saccharomyces cerevisiae. This work was further carried out by Cohen (1931) who connected 35 half-cell MFCs in series. Allen and Bennetto (1993) then utilized the understanding of electron transport chain to design a very Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00006-0 © 2020 Elsevier Inc. All rights reserved.

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Figure 6.1 Microbial fuel cell.

basic MFC. Kim et al., (1999) then explained the existence of electroactive species, which do not require a mediator molecule to transport electron to the electrode surface. Performance of an MFC is measured in terms of electricity generation and pollutant removal efficiency. The electricity generation is determined by measuring the potential difference between anode and cathode (voltage). Anode and cathode are usually connected with an external resistance, which acts as a load in the circuit. Current generation is measured by dividing the resistance from voltage. Power generated is calculated as a product of current and voltage. An important parameter called Coulombic efficiency is used to give a measure of the amount of electrons generated with respect to organic matter oxidized. This is an indirect measure of electron loss and is used to describe the efficiency of an MFC. Pollutant removal is measured in terms of organic removal and is denoted by the change in equivalent chemical oxygen demand (COD) between influent and effluent. Huge difference between stoichiometric cell potential and actual cell potential result from potential losses caused due to factors called internal resistance. Internal resistances are considered to be connected in series with external resistance, hence

Pharmaceutical wastewater treatment in microbial fuel cell

137

Figure 6.2 Microbial electrolysis cell.

effecting the power generation. Internal resistances are classified into activation losses, Ohmic losses, mass transport losses and internal currents. Activation losses occur due to electron losses while initiating the oxidation and reduction reactions at the electrodes. Ohmic resistances are caused during transport of electrons and ions in the solution, through the separators and from electrode to connecting wires. Mass transport losses occur while transport of substrate, electrons and oxygen to and from electrode. Finally, internal current are observed in the case of substrate crossing over across the membrane. Internal resistance and its various components can be determined by doing a polarization study. The various portions of polarization curve denote various resistances. Thus reactors should always be designed keeping in mind minimization of internal resistance. On the basis of the presence of membranes, MFCs can be classified into single chamber (Huang et al., 2017) or double chamber (Erable et al., 2009). And, on the basis of mode of aeration, MFCs can be classified into air cathode (Liu and Logan, 2004) and aqueous cathode (Utomo et al., 2017). Other than these, upflow (Lay et al., 2015), downflow (Zhu et al., 2011), miniature (Chang et al., 2017),

138

Integrated Microbial Fuel Cells for Wastewater Treatment

large scale (Lu et al., 2017), stacked (Yazdi et al., 2015), (Oh and Logan, 2006), salt bridge (Min et al., 2005), tubular (Tremouli et al., 2018) are the various other configurations used till date. Other than wastewater treatment and bioelectricity generation, other various applications of MFCs are powering underground monitoring devices (sediment fuel cells), powering up devices in remote locations, biochemical oxygen demand (BOD) sensors, body fluid batteries, desalination (MDC) and treatment of toxic pollutants or bioremediation, pollutant treatment being the widely explored area.

6.2

Application to pharmaceutical wastewater treatment

The presence of pharmaceuticals in water arises from two different sources: production processes of the pharmaceutical industries and the daily uses of pharmaceuticals resulting in their presence in domestic wastewater (Cecconet et al., 2017; Gadipelly et al., 2014). Other major point sources of pharmaceutically active compounds (PhACs) are wastewater treatment plants (WWTPs), which are not specifically designed for the removal of these compounds (Cecconet et al., 2017). The presence of PhACs in the environment and their long-term exposure in low concentrations and complex mixtures may result in acute and chronic damages, behavioral changes, and accumulation in tissues, reproductive damage, and inhibition of cell proliferation (Patneedi and Prasadu, 2015). Antibiotics, which most of the times reach the sewage system unchanged, result in the development of antibiotic resistance in the microorganisms of the aquatic bodies, changing the microbial community structure. Their effects get further heightened in synergy with other chemical components of the effluent. Diclofenac, a nonsteroidal antiinflammatory drug (NSAID) was reported to severely affect the population of Nepalese and Himalayan vulture, as a result, it was banned in 2006 (Oaks et al., 2004). Carbamazepine, another pharmaceutical compound, was seen to accelerate fish embryonic development (Qiang et al., 2016). Plants were seen to accumulate many drugs present in irrigation water, providing a potential pathway of exposure to humans (Prosser and Sibley, 2015). The various endocrine disrupting compounds such as bisphenol and ethinylestradiol are known to cause cancers, reproductive disorders, behavioral and learning problems, asthma, obesity, diabetes, feminization of fishes, and altered oogenesis in female fish populations (Rochester, 2013; Bergman et al., 2013, Huber et al., 2005). The various treatment techniques of pharmaceutical wastewater (PW) as listed in Table 6.1 require large area, energy input, and costly chemicals and generate toxic by-products increasing the cost of treatment to a great extent. A technology such as MFC can provide an efficient alternative as it has the capability of generating energy, do not require costly chemicals, and neither generates toxic by-products. PhACs present in trace concentration in wastewater are usually removed by mechanism of sorption or biodegradation. Ionic charges of the PhACs play a huge role in sorption, for example, positively charged compounds have shown better

Table 6.1 Various treatment techniques of pharmaceutical wastewater. No. Treatment type 1.

Process name

Physicochemical Nanofiltration and ultrafiltration

2. 3.

Drug name

References

Tylosin, amoxicillin, alkaline protease, EDCs

Sun et al. (2000), Zhang et al. (2003), Shahtalebi et al. (2011), Bezawada et al. (2011), Yoon et al. (2007) Alonso et al. (2018) Mestre et al. (2007), Yoon et al. (2003), Cyr et al. (2002), Goel et al. (2005), Li et al. (2016) Dodd and Huang (2007) Suarez et al. (2010), Go¨bel et al. (2005), Kruglova et al. (2014), Drillia et al. (2005) Villar-Navarro et al. (2018) Prasad et al. (2011), Wang and Gunsch (2011), London˜o and Pen˜uela (2015), Katsou et al. (2016), Hasan et al. (2016), Stadler et al. (2015), Vergili et al. (2018) Casas et al. (2015) Tambosi et al. (2010), Dialynas and Diamadopoulos (2012), Chang et al. (2008), Clara et al. (2005), Sipma et al. (2010), Radjenovic et al. (2007), Kimura et al. (2005), Urase et al. (2005), Kovalova et al. (2012), Hai et al. (2011)

Reverse osmosis Adsorption

Ciprofloxacin Ibuprofen, estrogens, mercury, lead, ranitidine

Chlorination ASP

Trimethoprim PPCPs, ibuprofen, carbamazepine, diclofenac, sulfamethoxazole

6. 7.

HRAP SBR

8. 9.

MBBR MBR

PhACs Ketoprofen, naproxen, carbamazepine, gemfibrozil ibuprofen, methylparaben, paracetamol, doxycycline, tetracycline, ibuprofen, amoxicillin PW Acetaminophen, ketoprofen, naproxen, carbamazepine, 17β-estradiol, 17α ethinylestradiol, chlofibric acid PhACs, fragrances, EDCs, diclofenac, ibuprofen, mefenamic acid, estrogens, sulfamethoxazole, real PW

4. 5.

Biological treatment

(Continued)

Table 6.1 (Continued) No. Treatment type

Process name

Drug name

References

10.

MBR embedded with activated carbon/silver NPs Anaerobic MBR Hybrid MBR with enzymes AD UASB

Sulfamethoxazole, carbamazepine

Li et al. (2011), Behboudi et al. (2018)

β-Lactam antibiotics PhACs EDCs and NSAIDs Tylosin, avilamycin

Svojitka et al. (2017), Huang et al. (2018) Ba et al. (2018) Samaras et al. (2014) Chelliapan et al. (2006), Sreekanth et al. (2009) Narumiya et al. (2013) Lucas et al. (2018), Badia-Fabregat et al. (2015) Naghdi et al. (2018) Afzal et al. (2007)

11. 12. 13. 14. 15. 16. 17. 18.

19. 20. 21. 22. 23. 24.

25.

AOP

Anaerobic sludge digestion Fungal degradation, Trametes versicolor (fungi) Fungal oxidoreductase enzymes Pseudomonas aeruginosa and Pseudomonas pseudomallei (bacteria) Chemical and electrochemical

Sewage sludge Carbamazepine, diclofenac, iopromide, venlafaxine Pharmaceuticals Phenol

Solar/TiO2/H2O2 Solar/TiO2/photo-Fenton Membrane nanofiltration and solar photo-Fenton Ozonation Electrocoagulation, photoelectrocoagulation, peroxicoagulation, peroxiphotoelectrocoagulation Sonoelectrochemical catalytic oxidation

Phenolic wastewater Acetaminophen, β-blocker atenolol Sulfamethoxazole, ibuprofen, ofloxacin, carbamazepine, flumequine PhACs PW, Ayurveda PW

Trimethoprim

Cephalosporin

Moreira et al. (2014, 2016), Domı´nguez et al. (2012) Adishkumar and Kanmani (2010) Radjenovi´c et al. (2009) Miralles-Cuevas et al. (2015) Huber et al. (2005) Farhadi et al. (2012), Singh et al. (2016)

Yang et al. (2016a,b)

26. 27.

28. 29. 30. 31. 32. 33. 34. 35.

Hybrid

Aerobic oxidation and chemical coagulation Combined Anaerobic/Microaerobic and two-stage biological process Anaerobic digestion and MBR TiO2 photocatalysis/RBC Fenton oxidation and aerobic degradation Fenton oxidation and ASP Ozonation and anaerobic MBR Photocatalysis and photo-Fenton (Fe-TiO2) Electrocoagulation and photocatalysis Ultrafiltration, reverse osmosis and electrochemical oxidation

PW

Raj and Anjaneyulu (2005)

Chemical synthesis based PW

Chen et al. (2011)

Chemical synthesis based PW Real PW PW

Chen et al. (2008) Talwar et al. (2018) Tekin et al. (2006)

PhACs Etodolac Real PW

Badawy et al. (2009) Kaya et al. (2017) Bansal et al. (2018)

PhACs

Boroski et al. (2009)

PhACs

Urtiaga et al. (2013)

AD, Anaerobic digester; AOP, advanced oxidation processes; ASP, activated sludge process; EDCs, endocrine disrupting compounds; HRAP, high-rate algal pond; MBBR, moving bed biofilm reactor; NSAID, nonsteroidal antiinflammatory drug; PhACs, pharmaceutically active compounds; PPCPs, pharmaceuticals and personal care products; PW, pharmaceutical wastewater; SBRs, sequential batch reactors.

142

Integrated Microbial Fuel Cells for Wastewater Treatment

removal than negatively charged compounds because of negatively charged bacterial biofilm (Wang et al., 2015). The combined oxidation/reduction environment in BESs provides a suitable environment for the removal of a variety of trace organic compounds such as pharmaceuticals. Since the degradation of different drugs has different mechanisms making the processes very specific, combination of different conditions and configurations such as acidic anolyte, biocathode, adjustment of organic loading, preacclimatization of biocatalysts, anode catalyst, physicochemical pretreatment, and electric stimulation might increase the treatability of a real PW in BESs (Table 6.2). Treatment of wastewater generated from the pharmaceutical industries is a very challenging task because of its recalcitrant nature, varying composition (as shown in Table 6.2), high toxicity, and low biodegradability (Ismail and Habeeb, 2017). These characteristics of PW make it difficult to be treated by conventional biological treatment processes due to inhibition of biocatalytic activity. Velvizhi and Mohan (2011) observed a 33% enhancement of catalytic activity of the microorganisms by providing an electrogenic environment in a bioelectrochemical treatment (BET) reactor as compared to a plain anaerobic reactor (AnT) treating real PW. The enhanced electron accepting conditions due to the presence of electrodes and complete neutralization of the reducing powers, electrons and protons by oxygen in the cathode chamber contributed to higher organic, phosphate, nitrates, turbidity, color, solids degradation in BET. Even at higher organic loading the BET showed better performance than the AnT. The LC50 value of the effluent of BET was also more than that of AnT. The biocatalytic activity however was not shown to be enhanced by the presence of a biofilm bearer (Ismail and Habeeb, 2017). The catalytic activity of the microorganisms can further be increased by acclimating them to the toxic pollutants prior to MFC treatment. Prior acclimation of the bacteria induces genetic mutations, which make them resistant to the toxicants by expressing proteins responsible for their degradation. It also enhances selective growth of the resistant bacteria by inhibiting the nonresistant ones. Such kinds of observations were made while treating a sulfonamide antibiotic, namely, sulfamethoxazole (SMX) in an MFC. Acclimation of the biofilm showed to enhanced SMX degradation. The MFC was fed with different SMX concentration wastewater and the SMX removal rates were found to increase with increase in concentration of SMX. However, the voltage generation and the TOC removal efficiency were seen to get affected after a certain SMX concentration. The MFC was reported to remove more than 90% of the antibiotic, and biodegradation was seen to be the functional antibiotic removal mechanism instead of biosorption. An aromatic degrading bacteria, namely, Thauera was seen to be dominant at the anode (Miran et al., 2018). Similar kind of observation was made by Liu et al. (2012), where a steroidal drug production wastewater (SPW) fed MFC showed a very different kind of microbial population than acetate fed MFC. Application of catalysts at electrodes has lately gained huge interest because of their capability to accelerate reaction kinetics at the electrodes. While cathode catalysts can enhance reduction reaction rates at cathode, anode catalysts along with enhancing the oxidation rate, can play very significant role in degrading pollutants

Table 6.2 Pharmaceutical wastewater composition. Sl. No.

Country

Composition

References

BOD (mg/L)

COD (mg/L)

Solid (mg/L)

pH

Sulfate (mg/L)

Nitrate (mg/L)

40

118 (TSS)

7.27 9.2 5.3 8.2 6 7 6.6 6 2.07 6.6 9.4 7.3 8.4 6 4.5

138 2914

750 10,800 1036 6 216 100 6350 200 2650 6 1457

810 6720 17,630 5000 60,000 5096 6 698 800 11,800 1753 9703 6 4880

7.23 760 (TN) 7130 (TN) 560 980 (TN)

4000

12,000 29,000 12,000

12

1727

12,425

13

830

4800

14 15 16 17

5992 6 143 3500 6 500 162 6 27

12,378 6 533 7000 6 800 1080 6 87 8877

1 2 3 4 5 6 7 8 9 10 11

Iraq China

Pakistan Taiwan Brazil Egypt Spain India

COD, Chemical oxygen demand.

600 2000 (TSS) 60 360 17,251 6 18,384 (TDS) 133.3 6 171 (TSS) 5000 (TS) 5442 (TDS) 4400 (TDS) 2890 (TSS) 8090 (TS) 4780 (TS) 1600 (TDS) 3180 (TSS) 1320 (TDS) 620 (TSS) 35,886 6 854.3 (TS)

12,840 (TDS)

8.5 9.04 7.45

893.7 376.8 6 364.4

7.9 0.94 6 0.41

Ismail and Habeeb (2017) Liu et al. (2012) Yang et al. (2016a,b) Chen et al. (2008) Afzal et al. (2007) Chang et al. (2008) Boroski et al. (2009) Badawy et al. (2009)

29.5 56

974 400

Domı´nguez et al. (2012) Yeruva et al. (2016) Velvizhi and Mohan (2012)

1128 1627 2280 6 30

Talwar et al. (2018)

5.8

5.07

526

20

Bansal et al. (2018)

7.9 6 0.32 5.2 6 6.8 8 6 0.4 10.7

9000 6 316 2500 6 350 163 6 16 440

3200 6 141

Raj and Anjaneyulu (2005) Chelliapan and Sallis (2011) Adishkumar and Kanmani (2010) Rajkumar and Palanivelu (2004)

144

Integrated Microbial Fuel Cells for Wastewater Treatment

in anode chamber. Also, the microbial consortium can be selectively enriched by using specific catalysts. For example, metal or metal oxide catalysts at anode, due to change in electrochemical properties, have shown to influence the growth of specific bacterial community. Also, anode modification with metal or metal oxides increases the surface roughness enhancing bacterial attachment and electron transfer at electrode. In one such work diclofenac (DCF), ibuprofen (IBF) and carbamazepine (CBZ) removal efficiency was studied in MFCs with anodes modified with MnO2, Pd and Fe3O4. Modified anode MFCs showed better power density and drug removal efficiency than carbon black control anode. Also, they showed high electrocatalytic activity and low internal resistance than the anode. Anode modified with MnO2 and Fe3O4 showed an abundance of Geobacter because of its capability to directly use MnO2 and Fe3O4 as electron acceptor. Along with Geobacter, Sphaerochaeta was found to be dominant on electrode modified with Pd, which was reported to produce hydrogen in BESs and was responsible for dehalogenation of DCF in presence of hydrogen (Xu et al., 2018). Although, effluent released from MFCs contain very limited toxic by-products, very few drugs might get partially degraded leaving behind active intermediates. Thus ensuring that the effluent from reactors does not contain any such compound that might revert back to their active form is necessary. For example, the regeneration of sulfonamides from their N4-acetyl analogues to their activated forms is often seen and presents a major risk. Thus in a study related to removal of sulfonamides, degradation of their N4-acetyl metabolites in MFC was also checked. The reactor was fed with synthetic wastewater containing acetate and a mixture of sulfonamides and their N4-acetyl analogues. SMX like the previous research was completely removed and did not lead to pharmaceutical reactivation like what is seen in aerobic processes. Electrochemical performance of the biocatalysts was also reported to be not affected by the antibiotics. However, not all the sulfonamide antibiotics were completely removed during the process, thus showing substrate specific activity of the anodic biofilm (Harnisch et al., 2013). The low biodegradability of PW increases the internal resistance of the MFC caused due to low metabolic rate of the biocatalyst and lesser efficient electron transfer from biocatalyst to anode and finally to cathode. A polarization study of the MFC can be used to identify the various potential losses, namely, activation losses, Ohmic losses, and concentration losses. One such study was done by Velvizhi and Mohan (2012) where the performance of an MFC fed with real PW was studied with respect to varying organic loading rates. Increase in organic loading showed decrement in activation losses due to higher availability of substrate and electrons. Ohmic loss was also seen to decrease with increasing organic loading. However, the concentration losses were found to increase with increase in organic loading because of the large oxidative force developed at anode due to more number of electrons released. But they can be canceled out due to decrease in activation and Ohmic losses. Thus the overall electrogenic activity of an MFC fed with PW of low biodegradability can be enhanced by adjusting the organic loading in the wastewater. PhAcs present in PW have often showed to improve the performance of MFC. For example, an SPW-fed MFC showed high power production of 23 W/m3 and better

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COD, efficiency of 82% which was equivalent to acetate fed MFC. This performance was attributed to the specific microbial consortium growth in anode due to prior acclimation. The voltage output, however, decreased after certain SPW strength, as the microorganisms were considered to get affected due to high osmotic pressure exerted by high strength wastewater. Similar kinds of results were observed by Wen et al. (2011), where penicillin addition to a synthetic PW improved performance. An MFC fed with synthetic wastewater containing glucose 1 penicillin showed relatively better performance than MFCs fed with wastewater containing penicillin and glucose alone. Highest power production of 101.2 W/m3 was observed in MFC containing 1 g/L glucose 1 50 g/L penicillin, which was six times higher than the sum of power production by MFC fed with glucose and penicillin as sole fuel. The current density was found to be 3.5 times higher in the case of MFC fed with penicillin than without penicillin. This better performance was attributed to the cell membrane permeability caused due to antibacterial properties of penicillin which reduced the internal resistance and enhanced the electron and substrate transfer. Also, 98% penicillin removal was achieved in 24 hours in the MFC fed with penicillin and glucose. While some drugs present in PW can enhance the performance of an MFC, some PhACs can have negative effect on the performance of the MFC in terms of voltage generation. For example, antibiotics such as aureomycin, sulfadimidine, roxithromycin, and norfloxacin have reported to have negative effect on the performance of an MFC. In the case of an MFC fed with these four different antibiotics, sulfadimidine showed highest inhibition followed by aureomycin, roxithromycin, and norfloxacin. On the contrary, the COD removal efficiency was seen to increase with addition of antibiotics. Roxithromycin addition, however, showed decrease in total phosphorous removal, whereas aureomycin showed a negative effect on the removal of total nitrogen (Zhou et al., 2018). Tobramycin, another antibiotic, was not seen to affect the current generation when added in concentration of μg/L. The current generation response was, however, affected, when tobramycin was added in the range of 0.1 1.9 g/L (Wu et al., 2014). Few of the broad spectrum antibiotics such as chloramphenicol (CAP) when treated via conventional biological processes are not completely removed and release toxic intermediates in the effluent. The chlorine and nitro groups present in the structure of CAP makes it biotoxic and resistant to the conventional biological treatment processes. Biocatalyzed cathodic reduction was seen to enhance chlorine reduction and conversion of nitro to amino group. Also, a minimum amount of energy input was seen to improve their removal efficiencies in BETs. Wu et al. (2017) did a study on the combined effect of applied voltage and cathode material on the removal of CAP in the cathode chamber of a BET which was initially operated in MFC mode and finally operated in MEC mode. Copper foam cathode at an applied voltage of 0.5 V was seen to perform better than other electrode materials nickel foam and carbon rod, in removing CAP. Higher applied voltage was observed to improve the removal efficiency, due to higher circuit current and hydroxyl radicals, generated in the cathode chamber (Wu et al., 2017). This electrical stimulation was also seen to have a significant effect on the chloramphenicol resistant bacteria (CRB) and expression of chloramphenicol resistant genes (CRGs). The CRBs were found to be

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abundant in more negative cathode potential, and the expression of the CRGs was induced in a less negative cathode potential. Thus an intermediate cathode potential was seen to improve dechlorination of CAP and helped in maintaining optimum CRB population and CRG expression (Guo et al., 2017).

6.3

Integration of microbial fuel cell with other wastewater-treatment processes

Wastewater generated from the pharmaceutical industries varies not only in composition but also in quantity, depending on the product, production process, and raw materials used. Thus approaches such as conventional physicochemical treatments, biological treatments, advanced oxidation processes, or hybrid techniques are used many a times combiningly to treat PW (Gadipelly et al., 2014). In such a study a sequential treatment strategy was designed by merging sequential batch reactors (SBRs) with BETs. Real field PW was first fed into two SBRs operated in aerobic/ anoxic conditions (SBRae/SBRan), the effluents from the SBRs were then fed to two BETs. The effluent from SBRan showed better COD removal than SBRae because in aerobic processes, substrate are oxidized to their intermediates and the electrons generated are accepted by oxygen as acceptor limiting the reduction reactions. While in anoxic condition, the reaction intermediates formed are further simplified in reduction reactions. The microorganisms present shuttle between anoxic and aerobic metabolisms behaving facultatively and enhancing the performance SBRan (68%) over SBRae (60%). BET1 showed a COD removal efficiency of 75%, while BET2 showed an efficiency of 73%. The cumulative COD removal efficiency was 91% in SBRanBET1, while it was 89% in SBRae-BET2. Similar observations were made in the case of multipollutant removal, namely, sulfate nitrate and phosphate, where higher pollutant degradation happened in SBRan-BET1 (92%, 87%, and 89%) than SBRae-BET2 (68%, 64%, and 60%). Higher bioelectrogenic activity was also observed in BET2 in comparison to BET1due to availability of simpler substrates (Yeruva et al., 2016). Paracetamol (PAM), a commonly found drug in PW can be mineralized in the presence of hydroxyl radicals, which can be supplied from the classical Fenton reaction between Fe21 and H2O2. But it requires an input of external energy, which was supplied by MFC in a work done by Zhang et al. (2015). Fenton reactions were conducted in cathode chamber, and peak analysis was done after regular interval of addition of PAM using UV VIS spectroscopy. Larger differences in peak intensities in the case of closed circuit operation in comparison to open circuit operation suggest the input flux of electrons from anode helped in degradation of PAM. During the initial stages, bioelectrochemical reduction of PAM to p-aminophenol was seen, giving a by-product of acetic acid. The second suggested reaction was hydroxyl attack on the benzene ring of PAM resulting in conversion of p-aminophenol to p-nitrophenol. The p-nitrophenol was then further broken down to volatile fatty acid by subsequent addition of hydroxyl radicals. However, this process was found to be highly dependent on initial iron dosage, pH of catholyte and external resistance applied (Zhang et al., 2015).

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Stacked constructed wetlands (CW) were used to power a BES treating SMX by Zhang et al. (2017). The CW-MFCs were reported to produce a stable voltage of 0.84 1.01 V, which helped in the rapid degradation of SMX in MFCs. Thus it was concluded that the SMX degradation was a reductive process requiring electrons from anaerobic reactions. The SMX degradation was found to follow first order kinetics and the products can be used as energy, carbon or nitrogen for growth and reproduction. Biotransformation was the chief process of degradation along with biosorption. The amount of voltage applied greatly influenced the microbial community in the BES reactor.

6.4

Large-scale microbial fuel cell: potentials and challenges

One of the greatest challenges in the commercialization of MFCs lies in proper designing of scaled-up versions of laboratory scale reactors. Till date laboratory scale MFCs have shown quite impressive results in terms of pollutant removal and voltage generation. To replicate those results in large scale, lot of factors have to be considered, for example, configuration, mechanical strength of the separator, cost of electrode and choice of catholyte or mode of oxidant supply, energy and space requirement, and many more. Electrode and separator material, mode of aeration, and pumping of the wastewater should be chosen keeping in mind their cost and amount of energy consumption. PW as discussed earlier is very much diversified in nature. Wastewater released from different units of a pharmaceutical industry is very much different in composition, particularly in organic content. The organic content of wastewater released from chemical synthesis unit usually varies between 800 and 29,000 mg/L. Such high concentration, if converted into energy has the potential to power own treatment units. Thus installation of BESs in WWTPs can play a significant role in reducing the energy requirement by supplying some part of it. No scaled-up reactor have been reported to be used till date to treat PW, hence the following section on large-scale reactors do not focus on any particular type of wastewater. Scaled-up reactors were found to show low current and power densities due to high volumetric Ohmic resistance and inactive reactor volume (Ieropoulos et al., 2008; Clauwaert et al., 2008). Thus instead of using single tank large reactors, module MFCs can be designed, which can be stacked to produce high power and current densities. In one such work, Dong et al. (2015) designed a 90 L pilot scale MFC, which consisted of five modules. The reactor was fed with brewery wastewater and was used to power its own pumping unit alternately. When fed with raw wastewater, the reactor showed COD and suspended solid removal efficiency of 82% and 86% respectively. Similarly, Ge and He (2016) designed a 200 L MFC containing 96 modules and could achieve 75% COD, 90% suspended solids, and 68% ammonia nitrogen removal. Internal resistance being a serious performance limitation can be dealt with by lowering the distance between anode, cathode, and separators. Membrane electrode

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assemblies as used by Lefebvre et al. (2011) in a 2.9 L MFC can reduce the distance, hence reducing the internal resistance. Flat plate modules can also prove to be a very promising configuration to reduce the internal resistance. Also, increasing the electrode surface area can provide enough space for attachment and development of the bacterial biofilm. Granular-activated carbon packed bed can perform very well in that respect. A stacked MFC of 72 L capacity with granular-activated carbon packed bed electrodes was designed by Wu et al. (2016) which showed a very high power density of 51 W/m3 and COD removal efficiency of 97% owing to the configuration of reactors and electrodes. They then further scaled up by designing a 1000 L MFC, which consisted of 50 modules and were fed with wastewater from two different MWTP. The operation was conducted for 1 year, and an average COD removal efficiency of 70% 90% was noted. High power density of 125 W/m3 was obtained when the reactor was fed with synthetic wastewater, while it decreased to 7 60 W/m3 when fed with real municipal wastewater. This shows lower biodegradability of real municipality wastewater (Liang et al., 2018). Separator as discussed before plays a huge role in contributing to the internal resistance of the MFC. Membrane less MFCs showed higher power and current density than MFCs with membranes. But an MFC without membrane has very little longevity and requires huge maintenance because of substrate diffusion to cathode and oxygen diffusion to anode. Such an attempt of designing large-scale membrane less MFC was made by Hiegemann et al. (2016), who installed a 45 L pilot scale MFC, which consisted of four single chamber membrane less MFCs. The reactor was installed downstream of a primary clarifier and COD, TSS, and nitrogen removal of 24%, 40%, and 28%, respectively, was observed. However, in another attempt made by Zhang et al. (2013), three tubular MFCs containing different ion exchange membranes and/or cathode catalysts, very dissatisfactory results were obtained which was totally not in agreement with similar kind of configurations and conditions used in lab scale reactors (Cha et al., 2010; Liu et al., 2011). Biofouling of cathode was considered to be reason of the failure.

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Oil and petrochemical industries wastewater treatment in bioelectrochemical systems

7

Surajbhan Sevda1,2, Vijay Kumar Garlapati3, Swati Sharma3, Udaratta Bhattacharjee4, Lalit Pandey1 and T.R. Sreekrishnan5 1 Department of Bioscience and Biotechnology, Indian Institute of Technology Guwahati, Guwahati, India, 2Department of Biotechnology, National Institute of Technology Warangal, Warangal, India, 3Department of Biotechnology and Bioinformatics, Jaypee University of Information Technology (JUIT), Waknaghat, India, 4Center for the Environment, Indian Institute of Technology Guwahati, Guwahati, India, 5Department of Biochemical Engineering and Biotechnology, Indian Institute of Technology Delhi, New Delhi, India

Chapter Outline 7.1 Introduction 157 7.2 Oil field and petrochemical wastewater treatment in the conventional treatment process 158 7.3 Oil field and petrochemical wastewater treatment in the bioelectrochemical system 164 7.4 Conclusion 168 References 169 Further reading 173

7.1

Introduction

Since the demand for energy resources is exponentially increasing, the concept of biofuel cells using microbes or enzymes as biocatalysts came into significance (Fig. 7.1). Here, we are discussing two categories of biofuel cells, namely, microbial fuel cells (MFCs) and enzymatic fuel cells (EFCs) which are well established for energy-saving waste disposal system. MFCs, as well as EFCs, are the new evolving technology not only for bioelectricity generation but also for wastewater treatments because of unique operating conditions (Sevda and Sreekrishnan, 2012, 2014; Quan et al., 2018). The current output in MFC is found to be low as compared to EFC. The conversion of fuel to electricity is more efficient in EFCs due to the absence of (1) cellular membrane, which inhibits electron transfer and (2) fuelconsuming microbial growth. But the commercialization of both the fuel cells still

Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00007-2 © 2020 Elsevier Inc. All rights reserved.

158

Integrated Microbial Fuel Cells for Wastewater Treatment

Figure 7.1 Diagrammatic depiction of a conventional fuel cell and a biofuel cell (Ivanov et al., 2010). A conventional fuel cell uses a metal catalyst in the anode, whereas biofuel cell uses microorganism or enzymes in the anode.

requires improvement for proper operations in the near future. Table 7.1 lists various MFCs and EFCs yield for various substrates. In MFC, microorganisms, which are biocatalyst, catabolize the substrate and therefore generate power. Carbohydrates, proteins, fatty acids can be used as fuel in this system. MFC can be categorized into (1) MFCs with the mediator and (2) MFCs without a mediator Current output in an MFC is attained by 1. 2. 3. 4.

catabolic activities of the microbes, oxidation occurring at the anode, reduction occurring at the cathode, and anode chamber to cathode chamber proton transfer (Chaturvedi and Verma, 2016).

Fig. 7.2 shows a flowchart of the current generation in MFCs and EFCs. MFCs can be used in the generation of electricity, production of biohydrogen, wastewater treatment, biosensor development for pollutant analysis, bioremediation of phenolic, and petroleum-based products (Rahimnejad et al., 2015; Sevda et al., 2016, 2017).

7.2

Oil field and petrochemical wastewater treatment in the conventional treatment process

Industrialization and urbanization simultaneously affect oil and petroleum resources worldwide. Petroleum and oil field resources generate a large amount of oily waste, which infects the water resources (Yu et al., 2017). The oil waste has a high amount of organic and inorganic pollutants, heavy oil complexes, surfactants and polymers, and their removal from the environment is the major concern facing by the oil refinery industries (Ji et al., 2002). The petroleum waste has aliphatic and aromatic hydrocarbon, and their presence enormously impacts every aspect of the

Table 7.1 Advantages and disadvantages of different treatment techniques for oil and petrochemical industries. Treatment

Advantages

Disadvantages

References

Rely on various physical, chemical parameters Reaction conditions vary with oil concentration

Rubio et al. (2002) Zeng et al. (2007)

Oil has a high amount of organic content, so it creates hurdles in flocculants formation Mainly depend on solids retention time

Zhang (2018)

Oil effluents treatment techniques Flotation Coagulation/ flocculation

Coagulation

Activated sludge process Chemical coagulation Micro-flocculation and filtration Electrocoagulation Advanced oxidation Microfiltration

A small amount of sludge formed Highly economical Decrease in oil concentration Elimination of suspended solids Poly-zinc silicates and anion polyacrylamide used as a coagulant Hydroxypropyl guar gum of oil field wastewater degraded Decrease in the COD Separate the petroleum hydrocarbon from the oil field wastewater BOD, COD, ammoniacal nitrogen, and oil concentration reduced Eliminate the suspended solids and oil Continuous flow of current density and consistent flow rate affect the turbidity and oil removal positively COD reduction Organic pollutants separated easily Excessive emulsifiers in oil can be segregated by the polytetrafluoroethylene membrane

Complex nature of effluent so alkaline, polymer flooding, and surfactant are used The little adsorption capacity of ceramsite filters Occupy huge space Homogenous electron Fenton does not show that much efficiency for oil removal and reduction in COD Emulsion concentration affects retention efficiency

Tellez et al. (2002) Yan et al. (2014) Si et al. (2018) Jiang et al. (2017) Boopathy and Das (2018) Kiss et al. (2013) (Continued)

Table 7.1 (Continued) Treatment

Advantages

Disadvantages

References

Different types of polymers used as a flocculent Microfiltration membrane not regenerated efficiently Parameters that affect the treatment process are not optimized At a certain temperature and pH condition, there is a decrease in the level of COD and sulfates removal This treatment proceeds in low temperature and pressure only

Zhong et al. (2003) Wang et al. (2009) Diya’uddeen et al. (2011) El-Naas et al. (2009)

Two different types of floatation cells are design for oil removal The necessity of optimization of EF and EC equipment in scaling-up at the industrial level

Santander et al. (2011) Dimoglo et al. (2004)

Only the Fe21 ions are effective for removal at high pH

Alta¸s and Bu¨yu¨kgu¨ngo¨r (2008) El-Naas et al. (2010)

Petroleum effluent treatment techniques Flocculation Microfiltration Photocatalytic degradation Electrocoagulation

Wet air oxidation

Flotation Electro Flotation/electro coagulation Precipitation and coagulation Adsorption

Decrease in the value of COD Oil removal increases Elimination of emulsified oil Mineralizes the inorganic and organic effluent Productive in nature Electrolytes efficiently remove the sulfate and COD It is mainly used for the pretreatment of highly contaminated wastewater BOD/COD ratio degradability enhance significantly Increase in COD removal Microwave-assisted CWAO efficiently separates the nonbiodegradable substances Efficiently eliminate the oil from the petroleum effluent Modified jet cell substantially affects the oily effluent There is no use of chemical reagents for the removal of oil, grease, and other contaminants Also, remove the suspended solids efficiently Iron as coagulant and calcium hydroxide as precipitant used for removal of contaminants Increase in the removal of sulfide and COD Date—pit AC act as the best absorbent for reducing the COD of petroleum effluent Less operating cost

Only batch adsorptions are assessed

AC, Activated carbon; BOD, biochemical oxygen demand; CWAO, catalytic wet air oxidation; EC, electrocoagulation; EF, electroflotation.

Sun et al. (2008)

Oil and petrochemical industries wastewater treatment in bioelectrochemical systems

161

Microbes or enzymes utilizing substrates acting as an electron donor

Electron flow

Migrate to anode

Migrate via electrical circuit to cathode

Proton flow

Migrate via electrolyte

Migrate via cationic membrane

Generation of electron and proton

Consumption via cathode and reduction of electron acceptor

Bioelectricity generation

Figure 7.2 Flowchart representing the pathway of MFCs for electricity generation. MFCs, Microbial fuel cells.

environment, human health, animals, and plant’s existence (Demirci et al., 1998). The dispose of this waste into the water or other streams requires treatments that modulate by the particular agencies and also not influence the local environment (Ghorbanian et al., 2014). The various treatments involved in oil refinery wastewater include filtration, coagulation, reverse osmosis, flocculation, and adsorption, mainly reducing the chemical oxygen demand. The economical and efficient method for the removal of oil and reduction in chemical oxygen demand (COD) makes these treatments most prominent (Do, 1998). The conventional treatment techniques which are in practice in handling the petrochemical effluents and oil filed effluents were depicted in Figs. 7.3 and 7.4, respectively.

162

Integrated Microbial Fuel Cells for Wastewater Treatment

Coagulation Photo catalytic degradation

Adsorption

Flocculation

Petrol chemical waste water treatment

Microfiltration

Wet air oxidation Precipitation

Activated sludge process

Figure 7.3 Conventional treatment processes for petrochemical effluents.

Flotation

Coagulation

Oil field waste water treatment

Flocculation Activated sludge process Advanced oxidation Microfiltration

Figure 7.4 Overview of the conventional treatment processes for oil field effluents.

The complex nature of the components in effluents of oil field and petroleum has not been eliminated by the treatment of the single method. There are certain types of methods executed by combining two treatment processes for improvement in the removal of contaminants efficiently and effectively (Santos et al., 2006). The operations of coagulation remove the suspended solids and colloids particles from the oil; these particles are not settled in the sedimentation and filtration process. Various types of coagulants and flocculants such as ferric chloride and aluminum sulfate are used for enhancement of COD and TSS removal in optimum conditions of pH, coagulant concentration (Altaher et al., 2011). The advantages and disadvantages of different treatment techniques have been summarized in Table 7.1.

Oil and petrochemical industries wastewater treatment in bioelectrochemical systems

163

Mollah et al. (2001) investigated the electrocoagulation treatments that are emerging technology where the electric field applied for the settling of colloidal particles effectively and their removal of turbidity and large or complex aggregates of oil effluent. The process of electrocoagulation did not involve chemical coagulant and pH maintenance. There are new types of electrocoagulation apparatus that has a continuous flow and mainly administer on the oil water emulsion. The inclined tubes used in this apparatus have a high efficiency of deposition and high oil removal rate (Jiang et al., 2017). The flocculation and flotation are the multistep systems for petroleum wastewater treatment. The petroleum has diverse toxic compounds and a high amount of oil sludge production, which are not treated by the simple methods (Kriipsalu et al., 2008). Petroleum has a mixture of different compounds such as sulfide, metallic, nitrogen, and oxygen. The concentration of sulfide varies from 1 to 150 mg/L, which is more than the desirable, permissible limits (Speight, 2014). The removal of sulfide by the chemical precipitation treatment from oil and petrochemical industries wastewater occurred using coagulants of ferrous sulfate and ferric chloride. There is another form of iron that eliminates the sulfur and reduces the COD (Alta¸s and Bu¨yu¨kgu¨ngo¨r, 2008). According to Li et al. (2006), there are various types of membrane, which is feasible and efficient for ultrafiltration treatment. The membrane adaptation by the organic and inorganic complex makes them more suitable for oil and COD reduction. Membrane moderation has antifouling execution and increases in flux recovery. The effluent generated from the oil industries are mainly in the form of oil in water emulsion. The floatation is the process that separates the oil by altering the density of two phases. There are new developments in the floatation treatment, which are induced and dissolved air flotation (Aurelle, 1985; Rubio et al., 2002). The treatment efficiency of these processes relates to the bubble size, velocity, and its gradient (Painmanakul et al., 2010). The oily effluents are the most toxic waste released by most of the refinery and oil industries. The methods of removal of these contaminants are different from industry to industry according to their needs and efficiency. The techniques generally involve in the pretreatment are -redox reaction, precipitation (Mbamba et al., 2015), and membrane filtration (Padaki et al., 2015). The micro-flocculation is the promising treatment for filtering the small suspended particles using polyaluminum chloride as coagulant and ceramsite as a filter medium (Si et al., 2018). The highly toxic nature of the petroleum effluent makes them more prominent against the release directly into the water streams. There are many types of traditional treatments that have many disadvantages, from sensitivity to the efficiency (Jung et al., 2015). Therefore from the past many years, the treatment processes are enormously changed by combining the two procedures for efficient removal of contaminants. The combine coagulation adsorption methods effectively remove the organic matter, other toxic chemicals, and reduce the COD (Wang et al., 2017). The wet oxidation is a simple conventional treatment by oxidizing the organic matter into the CO2 and H2O at an elevated temperature and pressure (Luck, 1999). The chemical oxygen demand of some effluents is very high due to the presence of

164

Integrated Microbial Fuel Cells for Wastewater Treatment

heavy contaminants. The heavy pollutants are not removed by the simple conventional methods, so microwave-assisted catalytic wet air oxidation treatment is an advance method having the property of eliminating nondegradable substances in optimized conditions of temperature and pressure (Sun et al., 2008). The supercritical water-oxidation treatment involves oxygen as H2O2 as an oxidant for the elimination of organic matter in oil effluent at standard temp and pressure condition (Kritzer and Dinjus, 2001). Ma et al. (2009) refer that conventional activated sludge treatment also has an excellent impact on refinery wastewater by reducing COD and other contaminants such as aldehyde, alkenes, phenol, and organic acids. The effluent of oil refinery contains phenol, benzene, xylene, toluene, and ethylbenzene (Akhtar et al., 2007) at very high concentration and their removal by the physical and chemical adsorption. The organoclay is an excellent absorbent for most of the BTEX compounds in their optimized condition of temperature, pH (Cavalcanti et al., 2012). The water recovery from petroleum wastewater generally performed after the conventional treatments on a pilot scale. RO technology is the most favorable and economical process for the use of that wastewater, which has a damaging effect on the environment. The petroleum industries are adopting electrodialysis reversal with reverse osmosis for removal of all chlorides and alkalies from the effluents (Venzke et al., 2018). The recent research attempts of different researchers towards the treatment of oil field and petrochemical effluents were included in Table 7.2.

7.3

Oil field and petrochemical wastewater treatment in the bioelectrochemical system

Wastewater from various sources such as petroleum wastewater, domestic wastewater, as well as other industrial wastewater such as a brewery and dairy are used as a substrate for power generation (Sevda and Sreekrishnan, 2014; Sevda and Abu-Reesh, 2017; Sevda et al., 2018). Using petroleum-based wastewater, the best results in terms of power output were found to be 225 6 1.4 mW/m2 and 156 mA/m2 (Table 7.3). However, these yields are not satisfactory as a real prospective capability of an MFC in terms of power requirements and therefore need appropriate modification of the operational units. MFCs, through efficient technology, have very few practical applications. This is because of the technical issues as well as scaling up of various factors that are yet to be solved related to its performance. First, high costs of the cation exchange membrane limit its use (Shantaram et al., 2005; Sevda and Sreekrishnan, 2012; Sevda and Abu-Reesh, 2018a; Sharma and Li, 2010). MFCs can be employed only in small devices such as biosensor because it faces low power and current output. The electrode’s surface area can be increased, and thus this problem can be overcome and improved the performance to a greater extent. Another remedy is to use a suitable power-management program by using ultracapacitors (Shantaram et al., 2005; Sevda et al., 2018; Sharma and Li, 2010). The commencement time for

Table 7.2 Research attempts in handling the oil field and petrochemical effluents. Types of effluent

Treatment/process condition

Output

References

Petroleum wastewater

Adsorption coagulation/pH: 3 11; stirring speed: 200 rpm; time: 30 min

Wang et al. (2017)

Petroleum effluent

Electro coagulation/Al electrode act as anode; time: 120 min; pH: 4 11; voltage: 12 V

Oily wastewater

Coagulation flocculation/Fe2(SO4)3 coagulant Polyelectrolyte Zetag8140—flocculants; pH: 3.5 4.5 Mixing: 50 120 rpm Time: 1 20 min Flocculation/microwave reactor with flocculants having capacity 400 W; time: 90 s; speed: 200 rpm; adsorption time: 30 min Coagulation microfiltration/ Coagulants: FeSO4
7H2O, Al2(SO4)3
18H2O FeCl2
4H2O and AlCl3
6H2O Membrane: mullite ceramic Time: 0 60 min Hermia’s model of filtration: for the study of permeation flux at different intervals of time Coagulation flocculation and flotation PAX 18, ferric sulfate: coagulant NALCO 71408: flocculant pH: 6 7

Oil and other pollutants removal efficiently executed by the PFMS and wooden AC Rise in deportation ability of the oil This treated wastewater enhances the thickness of plants Lowering of turbidity and increase in oil removal

Reduction in COD Microwave assistance improve ferric flocculants Process help refinement of wastewater An efficient procedure for the removal of contaminants

Cui et al. (2015)

Decrease in turbidity High removal of COD and total organic carbon

Santo et al. (2012)

Oil field wastewater Oily effluent

Petroleum wastewater

Sellami et al. (2016) Mousa and Hadi (2016)

Abbasi and Taheri (2013)

(Continued)

Table 7.2 (Continued) Types of effluent

Treatment/process condition

Output

References

Petroleum wastewater

Activated oxidation process/Fe21 and H2O2: Fenton reagent Time: 10 160 min pH: 7.6 UV, O3/UV, O3/H2O2, and TiO2: photo-Fenton reagent Chemical coagulation/ Temperature: 20 C 25 C pH: 5.5 8.5 Flotation/pH: 5 5.5 phosphoric acid and sulfuric acid for pH calibration Aeration rate: 4 6 ‘/min Impeller speed: 400 1100 rpm Ultrafiltration/Al2O3 PVDF membrane develop Temperature: 30 C Pressure: 0.1 MPa Crossflow velocity: 7.8 m/s

COD reduction

Tony et al. (2012)

It efficiently eliminates the suspended solids, color, and reduces the COD

Farajnezhad and Gharbani (2012) Welz et al. (2007)

Petroleum wastewater Oily wastewater

Oily wastewater

Aeration and speed of impeller affect the oil expulsion positively

Clarify the water from the oily effluent COD and TOC confinement up to 80% 90%

AC, Activated carbon; PFMS, polymeric magnesium ferric sulfate; PVDF, polyvinylidene fluoride; TOC, total organic carbon.

Li et al. (2006)

Table 7.3 List of different biofuel cells investigated using various waste substrates and their respective power density. Fuel type

Substrates

Microbes/enzymes

Power generation

References

Microbial fuel cell

Rapeseed straw hydrolysate Palladium nanoparticles Domestic wastewater 1 acetate Sewage wastewater Petrochemical wastewater

Wastewater treatment plant sludge

54 mW/m2

Jablonska et al. (2016)

Shewanella oneidensis Mixed exoelectrogenic microorganisms Anaerobic sludge Mixed microbes

499 6 11 mW/m2 257 mW/m2

Quan et al. (2018) Stager et al. (2017)

13.5 µW/cm2 156 mA/m2

Mixed microbes

28.27 W/m3

Electroactive biofilm

132 mW/m2

Anaerobic sludge

131 mW/m2 and 543 mA/m2 225 6 1.4 mW/m2

Passos et al. (2016) Chaturvedi and Verma (2016) and Hassan et al. (2018) Mohanakrishna et al. (2018a,b,c) Mohanakrishna et al. (2018a,b,c) Sjo¨blom et al. (2017)

Microbial fuel cell Microbial fuel cell Microbial fuel cell Microbial fuel cell

Microbial fuel cell Single chamber microbial fuel cell Double chamber microbial fuel cell Microbial fuel cell

Petroleum refinery and whey wastewater Petroleum refinery wastewater Sweet sorghum stalks

Glucose biofuel cell Hybrid biofuel cell

Petroleum refinery wastewater Waste hydrolysate Ferrocene-mediated EFC for domestic wastewater Electrolytic method Electrolytic method

Bioelectrochemical system

Petroleum refinery wastewater 1 acetate

Enzymatic fuel cell Enzymatic fuel cell

EFC, Enzymatic fuel cell.

Mixed microbes (iron and sulfurreducing bacteria) Chlorella pyrenoidosa Glucose oxidase Glucose oxidase Microbial anode with Shewanella MR1 with enzymatic airbreathing cathode Electrode biofilms

Srikanth et al. (2016)

3637 mW/cm2 200 nW/cm2

Lee et al. (2018) Kilic¸ et al. (2014)

290 µWh (1.04 J) 26 W/m3

Abreu et al. (2018) Higgins et al. (2011)

222 mW/m2 and 278 mA/m2

Mohanakrishna et al. (2018a,b,c)

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MFCs provides another shortcoming, which differs when subjected to change in various factors like construction design of the reactor, inoculums and electrode materials, operational factors such as substrates, temperature, exterior loading, etc. Scaling-up is also challenging because of the use of platinum as a metal catalyst, which is rather expensive. The practice of using open-air biocathodes and nonplatinized cathodes such as manganese dioxide may overcome these technical issues. Various cheap substrates, such as wastewater from different sources, can be utilized. The residual organic material in the waste can be used for the production of biopolymers such as polyhydroxybutyrate. Many genetic engineering approaches are applied for improving the metabolic pathway of the microorganisms, though it needs further research designs. EFCs have a substantial role, such as MFCs. It employs enzymes. But to improve and stabilize the power output of assembled cells is still a demanding mission, and at the same time organizing the systems to provide enough cell voltage for microelectronic devices power generation is still an ongoing effort by many researchers. EFCs typically work by cell loading to deliver satisfactory voltage for electronic devices (enzymes as catalysts in fuel cells) power output, which is relevantly better than conventional catalysts (Sakai et al., 2009). The most frequently used selective glucose enzymes are glucose oxidase (GOx) and pyrroloquinoline quinone glucose dehydrogenase. There are reports of EFCs that are combined with conventional fuel cells to enhance and improve performances because EFCs cannot alone generate adequate yield. Researches on EFCs have been ongoing since a quite long time for implantable bioelectronic devices so that to decrease our dependence on batteries (Ivanov et al., 2010). These fuel cells are capable of producing power density of 3637 mW/cm2 and 200 nW/cm2 using waste hydrolysate and domestic wastewater as substrates, respectively. These outputs are, however, better than the MFCs, although they do not meet the expected power productivity (Table 7.3). To increase the throughput and shelf life of the EFCs, hydrogels and definite polymers have been integrated. For a variety of medical devices used in recent medicine, EFCs can be used as an implantable power source (Ivanov et al., 2010; Sevda and Sreekrishnan, 2014; Sevda Abu-Reesh, 2018b). If the electron transfer between the electrode surface and the enzymes are made better, it can be efficiently used for improved power generation. The current discussion describes the recent advances in the various biological fuel cells, their bottle-necks, which, with further investigations, can progress in their performance, reduce the cost, and successful implementation of technology in field applications for a sustainable future.

7.4

Conclusion

In recent years, bioelectrochemical systems have gained a lot of attention due to simultaneous wastewater treatment and electricity generation possibilities. Wastewater from various sources such as petroleum wastewater, domestic sewage,

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as well as other industrial wastewater such as a brewery and dairy is used as a substrate for power generation using bioelectrochemical systems. However, current yields are not satisfactory as a real prospective capability of bioelectrochemical systems in terms of power requirements. Therefore it needs appropriate modification of the operational units and further research.

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Mbamba, C.K., Batstone, D.J., Flores-Alsina, X., Tait, S., 2015. A generalised chemical precipitation modelling approach in wastewater treatment applied to calcite. Water Res. 68, 342 353. Mohanakrishna, G., Abu-Reesh, I.M., Al-Raoush, R.I., 2018a. Biological anodic oxidation and cathodic reduction reactions for improved bioelectrochemical treatment of petroleum refinery wastewater. J. Clean. Prod. 190, 44 52. Mohanakrishna, G., Abu-Reesh, I.M., Al-Raoush, R.I., He, Z., 2018b. Cylindrical graphite based microbial fuel cell for the treatment of industrial wastewaters and bioenergy generation. Bioresour. Technol. 247, 753 758. Mohanakrishna, G., Abu-Reesh, I.M., Kondaveeti, S., Al-Raoush, R.I., He, Z., 2018c. Enhanced treatment of petroleum refinery wastewater by short-term applied voltage in single chamber microbial fuel cell. Bioresour. Technol. 253, 16 21. Mollah, M.Y.A., Schennach, R., Parga, J.R., Cocke, D.L., 2001. Electrocoagulation (EC) science and applications. J. Hazard. Mater. 84, 29 41. Mousa, K.M., Hadi, H.J., 2016. Coagulation/Flocculation Process for Produced Water Treatment. ,http://inpressco.com/category/ijcet.. Padaki, M., Murali, R.S., Abdullah, M.S., Misdan, N., Moslehyani, A., Kassim, M.A., et al., 2015. Membrane technology enhancement in oil water separation. A review. Desalination 357, 197 207. Painmanakul, P., Sastaravet, P., Lersjintanakarn, S., Khaodhiar, S., 2010. Effect of bubble hydrodynamic and chemical dosage on treatment of oily wastewater by induced air flotation (IAF) process. Chem. Eng. Res. Des. 88 (5 6), 693 702. Passos, V.F., Aquino Neto, S., de Andrade, A.R., Reginatto, V., 2016. Energy generation in a microbial fuel cell using anaerobic sludge from a wastewater treatment plant. Sci. Agricola 73 (5), 424 428. Quan, X., Xu, H., Sun, B., Xiao, Z., 2018. Anode modification with palladium nanoparticles enhanced Evans Blue removal and power generation in microbial fuel cells. Int. Biodeterior. Biodegrad. 132, 94 101. Rahimnejad, M., Adhami, A., Darvari, S., Zirepour, A., Oh, S.-E., 2015. Microbial fuel cell as new technology for bioelectricity generation: a review. Alex. Eng. J. 54 (3), 745 756. Rubio, J., Souza, M.L., Smith, R.W., 2002. Overview of flotation as a wastewater treatment technique. Miner. Eng. 15 (3), 139 155. Sakai, H., Nakagawa, T., Tokita, Y., Hatazawa, T., Ikeda, T., Tsujimura, S., et al., 2009. A high-power glucose/oxygen biofuel cell operating under quiescent conditions. Energy Environ. Sci. 2 (1), 133 138. Santander, M., Rodrigues, R.T., Rubio, J., 2011. Modified jet flotation in oil (petroleum) emulsion/water separations. Colloids Surf. A: Physicochem. Eng. Asp. 375 (1 3), 237 244. Santos, M.R., Goulart, M.O., Tonholo, J., Zanta, C.L., 2006. The application of electrochemical technology to the remediation of oily wastewater. Chemosphere 64 (3), 393 399. Santo, C.E., Vilar, V.J., Botelho, C.M., Bhatnagar, A., Kumar, E., Boaventura, R.A., 2012. Optimization of coagulation flocculation and flotation parameters for the treatment of a petroleum refinery effluent from a Portuguese plant. Chem. Eng. J. 183, 117 123. Sellami, M.H., Loudiyi, K., Bellemharbet, K., Djabbour, N., 2016. Electro-coagulation treatment and de-oiling of wastewaters arising from petroleum industries. J. Pet. Environ. Biotechnol. 7 (4), 1000290 (290). Sevda, S., Sreekrishnan, T.R., 2012. Effect of salt concentration and mediators in salt bridge microbial fuel cell for electricity generation from synthetic wastewater. J. Environ. Sci. Health, A: Toxic/Hazard. Subst. Environ. Eng. 47, 878 886.

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Sevda, S., Sreekrishnan, T.R., 2014. Removal of organic matters and nitrogenous pollutants simultaneously from two different wastewaters using biocathode microbial fuel cell. J. Environ. Sci. Health, A: Toxic/Hazard. Subst. Environ. Eng. 49, 1265 1275. Sevda, S., Abu-Reesh, I.M., 2017. Energy production in microbial desalination cells and its effects on desalination. J. Energy Environ. Sustain. 3, 53 58. Sevda, S., Abu-Reesh, I.M., 2018a. Improved salt removal and power generation in a cascade of two hydraulically connected flow-microbial desalination cells. J. Environ. Sci. Health, A Toxic/Hazard. Subst. Environ. Eng. 53 (4), 326 327. Sevda, S., Abu-Reesh, I.M., 2018b. Effect of the organic load on salt removal efficiency of microbial desalination cell. Desalin. Water Treat. 108, 112 118. Sevda, S., Dominguez-Benetton, X., Graichen, F.H.M., Vanbroekhoven, K., Wever, H.D., Sreekrishnan, T.R., et al., 2016. Shift to continuous operation of an air-cathode microbial fuel cell long-running in fed-batch mode boosts power generation. Int. J. Green. Energy 13, 71 79. Sevda, S., Abu-Reesh, I.M., He, Z., 2017. Bioelectricity generation from treatment of petroleum refinery wastewater with simultaneous seawater desalination in microbial desalination cells. Energy Convers. Manage. 141, 101 107. Sevda, S., Sharma, S., Joshi, C., Pandey, L.M., Tyagi, N., Abu-Reesh, I.M., et al., 2018. Biofilm formation and electron transfer in bioelectrochemical system. Environ. Technol. Rev. 7, 220 234. Shantaram, A., Beyenal, H., Veluchamy, R.R.A., Lewandowski, Z., 2005. Wireless sensors powered by microbial fuel cells. Environ. Sci. Technol. 39 (13), 5037 5042. Sharma, Y., Li, B., 2010. The variation of power generation with organic substrates in single-chamber microbial fuel cells (SCMFCs). Bioresour. Technol. 101 (6), 1844 1850. Si, S., Yan, Z., Gong, Z., Liu, P., Zhang, Y., Xiang, Y., 2018. Pilot study of oilfield wastewater treatment by micro-flocculation filtration process. Water Sci. Technol. 77 (1), 101 107. Sjo¨blom, M., Matsakas, L., Krige, A., Rova, U., Christakopoulos, P., 2017. Direct electricity generation from sweet sorghum stalks and anaerobic sludge. Ind. Crop. Prod. 108, 505 511. Speight, J.G., 2014. The Chemistry and Technology of Petroleum. CRC Press. Srikanth, S., Kumar, M., Singh, D., Singh, M., Das, B., 2016. Electro-biocatalytic treatment of petroleum refinery wastewater using microbial fuel cell (MFC) in continuous mode operation. Bioresour. Technol. 221, 70 77. Sun, Y., Zhang, Y., Quan, X., 2008. Treatment of petroleum refinery wastewater by microwave-assisted catalytic wet air oxidation under low temperature and low pressure. Sep. Purif. Technol. 62 (3), 565 570. Stager, J.L., Zhang, X., Logan, B.E., 2017. Addition of acetate improves stability of power generation using microbial fuel cells treating domestic wastewater. Bioelectrochemistry 118, 154 160. Tellez, G.T., Nirmalakhandan, N., Gardea-Torresdey, J.L., 2002. Performance evaluation of an activated sludge system for removing petroleum hydrocarbons from oilfield produced water. Adv. Environ. Res. 6 (4), 455 470. Tony, M.A., Purcell, P.J., Zhao, Y., 2012. Oil refinery wastewater treatment using physicochemical, Fenton and photo-Fenton oxidation processes. J. Environ. Sci. Health, A 47 (3), 435 440. Venzke, C.D., Giacobbo, A., Ferreira, J.Z., Bernardes, A.M., Rodrigues, M.A.S., 2018. Increasing water recovery rate of membrane hybrid process on the petrochemical wastewater treatment. Process. Saf. Environ. Prot. 117, 152 158.

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Wang, Y., Chen, X., Zhang, J., Yin, J., Wang, H., 2009. Investigation of microfiltration for treatment of emulsified oily wastewater from the processing of petroleum products. Desalination 249 (3), 1223 1227. Wang, B., Shui, Y., Ren, H., He, M., 2017. Research of combined adsorption-coagulation process in treating petroleum refinery effluent. Environ. Technol. 38 (4), 456 466. Welz, M.L.S., Baloyi, N., Deglon, D.A., 2007. Oil removal from industrial wastewater using flotation in a mechanically agitated flotation cell. Water SA 33 (4). Yan, H.Y., Xiao, M., Zhang, Z.Z., Li, J.Q., Shi, B.Q., 2014. Remediation of oilfield wastewater produced from alkaline/surfactant/polymer flooding by using a combination of coagulation and bioaugmentation. Pet. Sci. Technol. 32 (13), 1521 1528. Yu, L., Han, M., He, F., 2017. A review of treating oily wastewater. Arab. J. Chem. 10, S1913 S1922. Zeng, Y., Yang, C., Zhang, J., Pu, W., 2007. Feasibility investigation of oily wastewater treatment by combination of zinc and PAM in coagulation/flocculation. J. Hazard. Mater. 147 (3), 991 996. Zhang, Z., 2018. Combined treatment of hydroxypropyl guar gum in oilfield fracturing wastewater by coagulation and the UV/H2O2/ferrioxalate complexes process. Water Sci. Technol. 77 (3), 565 575. Zhong, J., Sun, X., Wang, C., 2003. Treatment of oily wastewater produced from refinery processes using flocculation and ceramic membrane filtration. Sep. Purif. Technol. 32 (1 3), 93 98.

Further reading Aljuboury, D.A.D.A., Palaniandy, P., Abdul Aziz, H.B., Feroz, S., 2017. Treatment of petroleum wastewater by conventional and new technologies-a review. Global Nest J. 19 (3), 439 452. Sevda, S., Abu-Reesh, I.M., 2019. Improved petroleum refinery wastewater treatment and seawater desalination performance by combining osmotic microbial fuel cell and upflow microbial desalination cell. Environ. Technol. 40 (7), 888 895.

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Kiran Tota-Maharaj1,2 and Marika E. Kokko2 1 Civil & Environmental Engineering Cluster, Faculty of Environment and Technology, Centre for Water, Communities and Resilience & The International Water Security Network, University of the West of England, Bristol (UWE Bristol), Frenchay Campus, Bristol, United Kingdom, 2Bio and Circular Economy Research Group, Department of Chemistry and Bioengineering, Tampere University, Tampere, Finland

Chapter Outline 8.1 Introduction 175 8.1.1 Urban stormwater 177

8.2 Energy recovery and stormwater treatment efficiency with bioelectrochemical systems 179 8.2.1 Influencing factors for bioelectrochemical system

179

8.3 Economic and environmental considerations

185

8.3.1 Environmental economics 185 8.3.2 Environmental impact and life cycle analysis 186

8.4 Technical scales of bioelectrochemical systems 187 8.5 Outlook, challenges, and future perspectives 189 8.6 Conclusion 191 References 192 Further reading 197

8.1

Introduction

Climatic and environmental changes, population growth, unstainable sources of fossil fuel reserves, rapid urbanization, needs for water and energy security, as well as waste generation, are some of the inherent issues associated with a conventional linear economy today. Countries across the globe need to gradually transform from traditional fossil fuel based economies to circular bioeconomies for long-term sustainability. The principles of a bioeconomy are to enable economic growth decoupled from increasing greenhouse gas emissions as a major contributor to climate change (Chandra et al., 2018). In the face of global epidemic environmental challenges, countries as well as the wider society need to rethink strategies and processes toward sustainable bio-based sectors. Long-term goals for the transition to a global bioeconomy are required for both energy and water security, making more Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00008-4 © 2020 Elsevier Inc. All rights reserved.

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efficient uses of resources and the downstream production of novel bio-based materials and useful products (Chiranjeevi et al., 2018). Important sectors for delivering feedstock for a bio-based economy are water, wastewater, stormwater, and related waste industries. Across the world, it is widely acknowledged that two economic models currently coexist; (1) the emerging bio-based economy (see Fig. 8.1) and (2) the fossil fuel driven economy, which have been prevalent for decades (Kru¨ger et al., 2018). The opportunities for bio-based economies are far and wide, underlining the much needed paradigm shift toward sustainability in order to meet a country’s environmental challenges and long-term sustainable goals. Fig. 8.1 gives a brief summary of the use of waste or industrial by-products that can be used in a bioeconomy. The development and diffusion of new green technologies and eco-innovations for a bio-based economy and the development of holistic approaches for sustainability from bio-based products are key drivers required for this paradigm shift—from conventional fossil fuels to alternatives—to occur. Thus the environmental acceptability of the whole bio-based processes, bioproducts, and downstream industrial applications is one of the key aspects that need to be considered in future designs of a bioeconomy (Fig. 8.1). The bio-based economy sits at the intersections of many overlapping concepts, including sustainable development, a circular economy, and green technologies such as microbial fuel cells (MFCs), microbial electrolysis cells (MECs), and other bioelectrochemical systems (BESs) (Fig. 8.1). These intersections (waste bioprocesses bioproducts industry) play a fundamental role in steering the transition process away from conventional fossil fuels. Initiating a transition toward such systems based on biomasses and circularity principles is a combination of scaling-up emerging innovative technologies (MECs and MFCs) involving the emergence of new sets of relations at the industry and downstream sectors, increasing production level as much as possible for consumers (Bose et al., 2018a,b).

Figure 8.1 Prospects of bio-based production of commodities from renewable sources.

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Beneficial utilization of waste streams (wastewater or stormwater) is to convert them into sustainable, low-cost materials stemming from the treatment processes of water and wastewater discharges (Trowsdale and Simcock, 2011). This is hugely advantageous environmentally in reducing the volumes of waste (both solid and liquid) and reducing resource consumptions in terms of materials, water, and energy usage. The resulting water process treatments can have lower associated costs with a subsequent higher efficiency, making “new water” resources available whereby water treatment was not previously viable. A futuristic bioeconomy should not only be based on sustainable and renewable raw materials but also has the assurance of a circular utilization of resources, interlinking biological and chemical transformations with technologies and systems to store and utilize bioelectricity and bioenergy (Holtmann et al., 2014). The transition to a bioeconomy opens up real possibilities for producing renewable products with energy valorization (Tommasi et al., 2012) from established markets and also opening up exciting opportunities for new industries and bioproducts.

8.1.1 Urban stormwater It is well understood in environmental engineering that stormwater (water stemming from precipitation rainfall, snowmelt, and ice) generated in villages, towns, and cities (constructed and urbanized areas) results in waste and pollution if not treated and managed properly (Barbosa et al., 2012). Like wastewater, stormwater must be dealt with appropriate infrastructure, because if it is improperly treated, especially during rainy conditions, it can cause severe impacts on receiving waters (Gasperi et al., 2014; Isteniˇc et al., 2012). Stormwater presents unique characteristics when compared to domestic sewage, industrial, and commercial wastewater (Gasperi et al., 2014). Stormwater pollution stemming from industrial, agricultural, and domestic activities can significantly affect the availability of clean water resources for communities. It can wash out pollutants from industrial, agricultural, and domestic sites and, thus, requires treatment. Stormwater ever so often may be unfit for use in agricultural applications and human consumption due to the presence of naturally occurring compounds. Urban stormwater is often conveyed to municipal wastewater treatment plants. The use of stormwater treatment technologies aims to reduce the pollutant loads as well as excessive runoff that enters into drainage systems. Stormwater contains and transports a variety of pollutants, both organic and inorganic. These include solids (suspended solids), heavy metals, biodegradable organic matter, organic micropollutants, pathogenic microorganisms, and nutrients (Hvitved-Jacobsen et al., 2010). Biochemical oxygen demand (BOD5), nutrient concentrations, and microorganisms are lower in stormwater when compared to wastewater, but if not treated properly, it can create negative ecological and environmental impacts (Table 8.1). The capture, treatment, and recharge of stormwater runoff can augment the water supplies of water-scarce regions. In water-limited climates, water supply potential exists for large-scale stormwater treatment, harvesting, and recharge, such as community-scale and larger infrastructure projects. Improving urban stormwater

Table 8.1 Characterization of stormwater pollutants. Pollutant group

Water parameters and units

Sources

Environmental impact

Suspended solids

Total suspended solids (mg/L); turbidity (NTU)

Heavy metals

Cu, Zn, Cd, Pb, Ni, and Cr (mg/L)

Solids can accumulate within sewer systems and be discharged at varying times Heavy metals are relevant because of their toxicity in the environment. Examples: copper (Cu), zinc (Zn), cadmium (Cd), and Lead (Pb)

Biodegradable organic matter

Biochemical oxygen demand (BOD5) and COD (mg/L)

Organic micropollutants and microplastics Pathogenic microorganisms

PAHs, PCBs, MTBE, and endocrine disrupting chemicals (μg/L)

Road and pavement wear, construction sites, infrastructure works, atmospheric debris, and anthropogenic wastes Vehicle components and parts, wear from tires, fuel and lubricating oils, traffic signs, and metallic road structures. Industries contribute to high loads of heavy metals entering stormwater Vegetation from leaves, trees, and logs; animals such as dogs, cats, and birds (fecal contributions or decaying matter); agricultural or industrial discharges PAHs from incomplete fossil fuel combustion, abrasion of tires, road asphalt, and pavements. Phthalate esters (urban construction plastic materials) Contributions from poultry farms, domesticated animals (cats and dogs), and birds

Nutrients

Total coliforms, fecal coliforms, Escherichia coli (CFU/mL)

Nitrogen; total kjeldahl nitrogen, nitrates (NO3) nitrites (NO2); total phosphorus (T-P); orthophosphate (PO32 4 ) in organic soluble phosphorous; SRP (mg/L)

Fertilizers, pesticides, herbicides, and atmospheric fallout

Organic matter from stormwater is sometimes less biodegradable as it is dominated by plant-based materials. There are a few problems associated with high organic loadings from CSOs A large percentage of compounds are discharged to the environment in small concentrations Stormwater sources contain varying concentrations when compared to domestic wastewater contributions from CSOs Nutrients can cause eutrophication, water discoloration, odors, toxic releases, and overgrowth of plants, encouraging invasive species

COD, Chemical oxygen demand; CSOs, combined sewer overflows; MTBE, methyl tertiary butyl ether; PAHs, polycyclic aromatic hydrocarbons; PCBs, polychlorinated biphenyls; SRP, soluble reactive phosphorus. Source: From Isteniˇc, D., Arias, C.A., Vollertsen, J., Nielsen, A.H., Wium-Andersen, T., Hvitved-Jacobsen, T., et al., 2012. Improved urban stormwater treatment and pollutant removal pathways in amended wet detention ponds. J. Environ. Sci. Health, A 47 (10), 1466 1477 and Barbosa, A.E., Fernandes, J.N., David, L.M., 2012. Key issues for sustainable urban stormwater management. Water Res. 46 (20), 6787 6798.

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quality, particularly with respect to pathogens and organic contaminants, is required to better meeting global treatment requirements (Table 8.1). The multiple benefits of water supply, urban amenities, and pollution reduction are important aspects for stormwater treatment systems where the source control is challenging. Novel bioelectrochemical technologies or BESs such as MFCs or MECs have great promise for improved on-site operation and stormwater treatment but must be demonstrated in field trials. Strategies to deal with stormwater are needed at different scales and there are opportunities for bioprocesses such as MFCs or MECs in the process (Tommasi et al., 2012; Stager et al., 2017). The intermittent nature of stormwater runoff can impact the treatment performance of BESs such as MFCs or MECs (Jafary et al., 2018) positively.

8.2

Energy recovery and stormwater treatment efficiency with bioelectrochemical systems

There are many publications on bioelectrochemical water treatment systems. High organic removal rates ( . 90%) and wastewater treatment efficiencies have been reported by several researchers for BES (Cusick et al., 2010; Majumder et al., 2014; Kim et al., 2016; Xie et al., 2017). The treatment efficiency is a function of the biochemical (biological) oxygen demand and chemical oxygen demand (COD) removal (%) combined with the energy recovery potential measured as kWh/kg of COD or kWh/kg of BOD present in stormwater or wastewater. Cusick et al. (2010) found that energy recovery and organic removal from wastewater can be more effective with MFCs than MECs. However, hydrogen production from wastewaterfed MECs can also be more economical based on bioelectrical energy requirements. Furthermore, Kim et al. (2016) reported power densities of 700 750 mW/m2 using low-strength wastewater similar to that of stormwater with a COD of 7 8 g/L. Treatment of low-strength wastewaters using MFCs has been effective at varying hydraulic retention times (HRTs) similar to aerobic processes, but the treatment of high-strength wastewaters can require longer HRTs. The performance of MFCs or MECs hydraulically connected in series can treat high-strength wastewater or stormwater with a COD ranging from 6 to 9 g/L/day (Kim et al., 2016). Nevertheless, power generation by BES can decrease significantly by up to 85% due to cathode fouling after extensive use ( . 180 days of operation). Organic pollutant removal rates can be improved by utilizing lower external resistance in the cell (Kim et al., 2016; Koo´k et al., 2016; Xie et al., 2017; Shen et al., 2018).

8.2.1 Influencing factors for bioelectrochemical system The performance of BES for stormwater treatment depends upon various factors such as pH, conductivity, electrode materials, the surface area of electrodes, membranes, electrical resistance and ohmic losses within the system, and substrate-type concentration reduction (He et al., 2006). Table 8.2 presents physicochemical

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Table 8.2 Stormwater quality runoff in Caribbean Small Island Developing States (2015 18). Constituent and units (mg/L, unless noted) 

Temperature ( C) Turbidity (NTU) pH Hardness (as CaCO3) Dissolved oxygen BOD5 Total dissolved solids Total suspended solids Electroconductivity (μmhos/cm) Chlorides Fecal coliform bacteria (CFU/100 mL) Nitrate nitrogen Ammonia nitrogen Phosphates (as P) Arsenic (μg/L) Cadmium (μg/L) Chromium (μg/L) Copper (μg/L) Iron Lead (μg/L) Mercury (μg/L) Zinc (μg/L)

Mean

Standard deviation

22.8 19.5 7.3 53 8.6 110.4 108 7.9 321 10.8 354,000 1.3 0.66 0.87 , 0.5 , 10 , 5.5 , 0.9 0.35 , 0.25 , 0.32 27

5.7 13.2 0.5 21 2.1 22.8 36 9.5 647 11.7 4350 1.1 2.4 0.71 9.6 22 23 69 0.11 0.72 2.5 125

parameters from a study conducted in a Caribbean Small Island Development State, illustrating the average values and deviation of parameters found. The applied voltage to BES is theoretically equal to voltage plus overpotential (the excess required potential due to resistance), which includes an ohmic overpotential (ion transfer between electrodes), anode and cathode overpotential (due to metabolism and activation losses), as well as concentration overpotential (increasing cathode chamber pH). If energy needs to be applied, then the BES becomes an MFC whereby the cell voltage produced is decreased by overpotentials. The effect of a few key parameters influencing the bioenergy recovery and wastewater treatment of MECs and MFCs are discussed next.

8.2.1.1 pH The pH of the medium plays a major role in MFCs and MECs, as it governs both the thermodynamic and kinetic reactions. For each unit rise in pH the anodic potential can decrease by approximately 20.06 V (Nimje et al., 2011; Khan et al., 2012). For BESs using ion-exchange membranes, variation in pH is a common phenomenon that results in BES due to oxidation reactions decreasing the pH at the anode

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and reduction reactions, increasing the pH at the cathode. Liu et al. (2014) reported that pH 9 is optimum for H2 production and COD removal with MECs. Meanwhile, the gas production rate decreases at a pH value higher or lower than the optimum pH. However, this is not always the case with biological systems. An alkaline pH is always suitable for bacteria, while fungi require an acidic pH for its activity (Khan et al., 2014). Moreover, the exoelectrogenic population of microorganisms also requires a high pH for transferring their electrons to the anode (Nimje et al., 2011; Khan et al., 2012). On the other hand, the authors’ research has shown that the operation of a BES can occur at pH 1.5 with simultaneous electrical current production without the need for external mediators. Nimje et al. (2011) studied the effects of pH on the performance of MECs for wastewater treatment and observed an increase in electrochemical activity with an increase in pH of the anolyte from 7 to 9. Likewise, several other researchers have reported that an alkaline pH is more favorable to enhance bacterial activities at the anode (Koo´k et al., 2016; Xie et al., 2017). Further, Rozendal et al. (2008a) reported a loss of output cell potential by approximately 0.4 V, when pH is increased to around 6 U.

8.2.1.2 Temperature The temperature of stormwater likely varies depending on location as well as on the season. Furthermore, the performance of MFCs and MECs is affected by a change in temperature. For MECs the main part is an anaerobic (anodic chamber where the actual degradation of organics takes place), wherein the activity of the microorganism is temperature dependent (mesophilic phase occurs at 35 C 40 C) (Khan et al., 2014). The extent of variation of electrogenesis with BES is much lower than methanogenesis. Khan et al. (2014) studied the performance of MEC at three varying temperatures: 20 C, 25 C, and 30 C. The columbic efficiencies and COD removal rates peaked at 30 C. Moreover, BES such as MFCs and MECs can perform even at a lower temperature (#20 C) in comparison to many other biological reactors and digesters (Pham et al., 2006).

8.2.1.3 Electroconductivity The electroconductivity of stormwater can affect the performance of MFCs or MECs (Table 8.2). Catholyte and anolyte solutions of higher electroconductivity can enhance ion transfers that improve the overall performance of BES by decreasing the internal resistance of the cell (Logan et al., 2006; Hamelers et al., 2010; Rosenbaum et al., 2011; Ki et al., 2016). Moreover, Logan et al. (2006) reported that ohmic losses could be reduced by using solutions of higher electroconductivity. Call and Logan (2008) reported an increase in the H2 production rate of MEC when the electroconductivity was increased from 7.5 to 20 mS/cm. Increasing the electrode spacing from 0.4 to 1.4 cm and reducing solution conductivity from 7.8 to 1.8 mS/cm can upsurge the internal resistance of the system (Hutchinson et al., 2011).

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8.2.1.4 Kinetics and thermodynamics The thermodynamics and kinetics are important performance parameters of BES. Organic matter and microbes spontaneity in the system, as well as the feasibility of the process, are critical (Rozendal et al., 2008a,b; Rabaey and Rozendal, 2010), meaning the activity of the microbes and their affinity to the substrate. The spontaneity of biological and chemical reactions has been linked to the thermodynamic parameter ΔG, which is defined as the change in Gibbs free energy (Rozendal et al., 2008a,b; Rabaey and Rozendal, 2010; Khan et al., 2017). Analysis of the kinetics and thermodynamics of biological and chemical reactions is very critical, and only those reactions that are kinetically and thermodynamically favored will yield the expected products with BES. For a spontaneous reaction, ΔG should be negative, and in the case of MECs, ΔG is positive for most of the biochemical reactions (Rozendal et al., 2008a,b; Rabaey and Rozendal, 2010; Khan et al., 2017). ΔG of the whole BES cell is positive in this case, it is calculated from the Gibbs free energies of the individual anode and cathode reactions. An external energy source is required to drive these reactions with MECs. Since the microbial degradation of substrate is a spontaneous phenomenon (ΔG is negative for the oxidation reactions), the reaction starts as soon as the anodic chamber is fed with carbon sources resulting in the formation of electrons (e2) and protons (H1) (Rozendal et al., 2008a,b; Rabaey and Rozendal, 2010; Khan et al., 2017). A second reaction occurs (the reduction of protons) that is a nonspontaneous reaction in BES, since the fusion of two protons will require an external energy source (Logan et al., 2006; Hamelers et al., 2010; Rosenbaum et al., 2011; Ki et al., 2016). In MFCs the oxygen reduction is spontaneous, while in MECs, the reduction of protons to form H2 requires an external energy source.

8.2.1.5 Electron transfer mechanism Previously, it has been assumed that microorganisms transfer the electrons to electrodes in two ways, (1) directly through membrane-associated transfer or (2) indirectly using chemical mediators. Direct electron transfers are only possible when there is physical contact between the outer cell membrane and the surface of the electrode, for example, with outer membrane cytochromes or with nanowires (Holzman, 2005; Zhang et al., 2009). Some nanowires can work on the principle of superexchange electroconductivity. The nanowires are electrically conducting appendages that are produced by microorganisms to communicate with the external electron acceptors (Boesen and Nielsen, 2013). Various groups of bacteria have confirmed efficient electron transfer mechanisms, such as beta- and delta-proteobacteria within BES (Kiely et al., 2011). Otherwise, specific microbes can excrete chemicals that can act as redox compounds to perform indirect electron transfer (Huang et al., 2011). These electron donors can also be reduced products of microbial metabolism that can transfer the electrons to the anodic chamber (Mohan et al., 2014; Rosenbaum et al., 2011). The electron shuttle or mediators get reduced by accepting the electrons within the cell

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membrane, thereafter passing through the membrane and transferring its electrons to the anode in a cyclic fashion (Du et al., 2007).

8.2.1.6 Electrodes and applied potential Electrode material selection depends on bacteria and the reactions, which take place in the cell (Mohanakrishna et al., 2016). As metabolic reaction rates depend on carbon source and biofilm stability in the anode, the electrode material has a significant effect on cell performance. Generally, electrodes should have the following features: moderate-to-high electroconductivity; biological, chemical, and physical stability; economically worthwhile; and have a high specific surface area (Mohanakrishna et al., 2016). For anodes and cathodes, carbon-based materials can be used and available with a wide range of costs from the US $1000.00 to d10 50/1 m2 (Call et al., 2017). Graphite fiber brushes are a promising material for anodic reactions (Rozendal et al., 2006). It has been widely recognized that metallic electrodes with coated carbon are a good choice for BES due to higher conductivity and strength of metals, along with high biocompatibility of carbon-based materials. Copper and stainless steel are not suitable for anode material due to corrosion and toxicity. Moreover, the overpotential (energy lost at the electrode) and the coulombic efficiency (produced electrons originating from the substrate that ends up in the targeted product) indicate the efficiency of electrodes. Graphite electrodes can serve as an electron sink for anaerobic respiration (Gregory et al., 2004; Rosenbaum et al., 2011). This can have environmental benefits, for example, for bioremediation of stormwater. Carbon-based materials have been used broadly as electrodes. The conductivity of carbon-based electrodes is fairly low in comparison with metals; however, using metallic current collectors can compensate for this drawback (Gregory et al., 2004). Metallic electrodes (such as titanium and stainless steel) are not suitable for bioanodes due to a lower surface area and surface properties, which results in less biofilm development. When considering the costs and benefits of carbon-based materials, it is recommended to utilize these materials as electrodes in BES, especially in systems aiming at wastewater, such as stormwater treatment. Using graphite- and carbon-based materials as electrodes has various benefits, such as suitable cost, higher surface area (due to porosity), and stability, but they have lower mechanical strength compared to metallic electrodes (Yu et al., 2016; Borenstein et al., 2017).

8.2.1.7 Membranes The membrane is used to separate anode and cathode compartments and to enable ion exchange between the chambers. By using membranes, produced gases are isolated in each compartment, the microbial section is separated from the other section, and it leads to the prevention of contamination of the membrane (Krieg et al., 2018). For most BES, the use of a membrane is unavoidable, but the costs of the membrane are challenging, and finding alternative solutions is an area for future

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research. The operation of MECs and MFCs depends on membrane selectivity and performance, for example, affinity to anions or cations and the longevity of the membrane. Many considerations and selections must be taken into account for membrane selection and usage. Generally, controlling the operating conditions in the anode and cathode chambers, such as pH and concentration of substrate and O2, without avoiding mixing of the fluids, is the advantage of using membrane. However, the membrane increases the overpotential of the system (Krieg et al., 2014). Membranes are the most favorable separators for BES operating in temperatures between 30 C and 37 C and when the pH is around 6 7 (neutral pH) which are appropriate conditions for microbial metabolism (Krieg et al., 2018). The most common type of membrane is the proton exchange membrane (PEM), which allows the needed flow of protons through the membrane to supply the cathode section. Cation- and anion-exchange membranes are also common with BES. New membrane material, such as forward osmosis membranes, can increase the reactor performance for wastewater or stormwater treatment in MEC (Krieg et al., 2014, 2018). A number of materials have been suggested for membranes in BES, commonly polymer electrolyte membranes. Nafion as a PEM is the reference in most studies; however, its high cost is a drawback. The Nafion-117 PEMs (chemically stabilized perfluorosulfonic acid polymer) have been selected by most laboratory-scaled systems as the preferred membrane material for MECs and MFCs. Ultrafiltration membrane can be an alternative that is less expensive and has high ionic resistance (Zaidi, 2009). Other alternative membranes include Hyflons, Zirfons, and Ultrexs Cation Exchange Membrane (CMI) 7000 as PEMs, anion or cation exchange membranes. Sulfonated polymer membranes, such as sulfonated polyether ether ketone and disulfonated poly (arylene ether sulfone) membranes, as well as nanocomposite polymer membranes, can be used in BES to compensate for the high cost associated with Nafion, which is the preferred membrane choice for MFCs (Leong et al., 2013). Carbon cloth can also be used between cathode and anode chambers for BES (Yang et al., 2017). However, there are many challenges with carbon cloth due to biological degradation overtime. In terms of membrane structures, heterogeneous membranes have been reported to possess higher mechanical strength but have higher resistance, which results in greater ohmic overpotential (Logan, 2010). On the other hand, homogenous membranes are thinner and cause lower resistance (Ki et al., 2016). In some cases, one chamber is used (instead of using separated compartments) because membranes can cause a pH gradient in the reactor between the electrodes (Thygesen et al., 2010; Logan, 2010). In addition, membranes can lead to ohmic resistance and consequently increase the energy consumption within BES (Logan, 2010). The membrane itself causes overpotential in the system. Around 2400 to 21000 mV is required to compensate energy losses as a result of electrodes and ion-exchange membranes overpotentials (Pisciotta et al., 2012). Restriction of oxygen diffusion is one of the main reasons for using separated chambers. Oxygen diffused at the anode is able to consume electrons instead of their transfer to the anode electrode, leading to reactor efficiency reduction. In addition, some anodic bacteria are sensitive to oxygen.

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Nevertheless, an advantage of membrane-less single-chamber reactors includes reducing the cost, dipping potential losses caused by membranes, and enhancing energy recovery (Hashemi et al., 2012).

8.3

Economic and environmental considerations

Both MFCs and MECs can become practical and sustainable water, wastewater, and stormwater treatment technologies. However, critical issues still remain to be considered on the viable commercialization and technological feasibility of BES. The levels of performance for both MECs and MFCs in becoming an attractive technology from an economic perspective are still blurred. MECs and MFCs can become sustainable and practical water and wastewater treatment technologies once they fulfill the following criteria: (1) they are technically feasible and produce treated effluent that meets national and international environmental standards and (2) they are economically feasible for operation and maintenance (waterenergy footprint, ancillary equipment). While converting the organics in stormwater into electricity, some ions (nutrients, etc.) could be removed/recovered by using a 3-chamber set-up. Furthermore, these technologies must be environmentally viable and compatible with environmental protection policies and laws. Escapa et al. (2016) noted that these systems must meet requirements such as (1) technical feasibility (ability to produce water effluents that meets national and international legal and environmental standards for water quality with no impact on the aquatic environment); (2) economic viability (reduction in capital and maintenance cost, smaller water and land footprint, manufacturing, installation and ancillary requirement, reduction in operational costs with regards to energy consumption, chemical additions, and labor costs when compared to conventional treatment systems); and (3) environmental achievability (the systems must be compatible with the environment and the protection of natural habitats and in line with environmental policies).

8.3.1 Environmental economics MFCs and MECs are not industrial technologies yet. Although there has been a great deal of applied research, much of it is still applied primarily in academia and needs to be seen on large-scale applications. Full-scale MECs or MFCs producing bioenergy (hydrogen) and bioelectricity are technically feasible. Nonetheless, the cost of construction, the microbial fluctuations, the changes in physiochemical wastewater parameters and likely wastewater effluent standards, the power outputs, and the coulombic efficiencies of these systems have to be carefully considered. Currently, the capital cost exceeds that of anaerobic digestion (AD), sustainable urban drainage systems, or activated sludge process (Yu et al., 2018). From water to energy savings, BES be derived from MEC or MFC technologies is undoubtedly an economic benefit and financially appealing when compared to

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the conventional treatment works for wastewater or stormwater (Tota-Maharaj and Paul, 2015). Reports and studies by Cusick et al. (2010) have translated energy savings into financial terms and estimated costs. It was found that BES could be as high as US $0.20 kg/COD if prices of H2 remained stable at US $6/kg. Other industrial wastewater treatment flows from breweries and wineries using MEC systems computed net financial gains of US $0.06/kg of COD at a laboratory scale (Cusick et al., 2010). The benefits of BES, such as MECs, are the low production of biomass when compared to aerobic treatment processes (Pant et al., 2012). This is an important feature as the management of biomass accounts for a substantial fraction of the general operating cost of water and wastewater treatment works (Curtis, 2010). The energy consumption in wastewater treatment works currently accounts for up to 50% of its operating costs. Studies by Gandiglio et al. (2017) found that using MEC technologies as a pretreatment before aerobic treatment in wastewater treatment plants could reduce energy consumption by 20%, which may represent a significant reduction in the utility’s energy bill. Whilst these encouraging calculations are noteworthy, the huge capital and operating cost associated with BES are the main barriers to commercialization and industry use. Although there are many challenges to be solved with the development of both MFCs and MECs, researchers across the world are still progressing. Interestingly, it was reported by Escapa et al. (2016) that the feasibility criteria for MFCs were much more stringent than for MECs, suggesting that the prospects for large-scale MECs were better than that of MFCs. Previous research projects found that capital cost of approximately d100/m2 which includes electrodes, membranes and materials, and an operational cost of about d0.10/kg of COD removed equates to an electricity cost of almost d0.06 kW/h and a hydrogen gas price of d0.40/m3 (Escapa et al., 2016). Furthermore, Escapa et al. (2016) estimated that current densities greater than 5 Amps/m2 and energy consumption less than 1.0 kWh/kg of COD removal justify the use of MECs for hydrogen production to particular wastewater treatment plants, providing that the cost of future BES technologies does not exceed Bd1500/m3. Nevertheless, the cost of ancillary equipment such as hydrogen storage tanks, gas compressors, and rectifiers has not been taken into account as the direct cost associated with MECs when compared to conventional water and wastewater treatment.

8.3.2 Environmental impact and life cycle analysis Feasible BES designs need to compete with alternative stormwater and wastewater treatment systems first with the added benefit of energy production, not only economically or technically but also environmentally (Pant et al., 2012; Foulet et al., 2018). The life cycle assessment (LCA) is a globally standardized system for assessing and evaluating environmental impacts of projects, products, or technologies across its entire life cycle, from sourcing of materials to end of use and waste disposal (Foulet et al., 2018). More importantly, LCAs provide valuable information and useful methods for comparing various technologies such as AD or similar

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processes (e.g., membrane bioreactors) with regards to environmental factors, eco-toxicities, aquatic pollution (water acidification), air pollution, and ozone layer depletion, which are not usually considered in the field of BES. Previous research by Pant et al. (2011) looked at a compilation of inventory factors such as boundaries, inputs, outputs, and impact assessment, which are relevant to the LCA of MECs. Earlier research (Foley et al., 2010a,b) compared the environmental impact assessment of high-rate anaerobic treatment of wastewater to MFCs and hydrogen peroxide producing MECs. These studies concluded that MFCs did not differ significantly from conventional AD and MECs provided the environmental benefits, through the displacement of conventional chemical production (Foley et al., 2010a,b). In addition, BES systems such as MFCs and MECs are independent of variabilities from some of the less reliable and stable factors from a life cycle inventory such as construction materials and energy fluxes (Pant et al., 2012; Foulet et al., 2018). Further analysis of the environmental benefits of using MEC or MFC technologies to treat wastewaters, such as stormwater, is needed.

8.4

Technical scales of bioelectrochemical systems

There are many hurdles to overcome if BES is to offer a sustainable future in the water and wastewater sector. Most of the scientific outputs and research have been performed at a laboratory scale, using simple substrates, often at a controlled temperature and pH. Although of great value in improving our understanding of MECs, these studies do not tell us about the challenges or even benefits of running such systems at a larger scale, with real wastewaters in temperate climates. There is a greater need to demonstrate that these systems can work at a larger scale and under realistic conditions, elevating the technology from a laboratory curiosity into a practical solution to a global environmental problem (Montpart et al., 2015). Several studies have provided valuable insights into the operation of MECs and MFCs; however, it has not provided a proof of concept that real, unsupplemented wastewaters can be used to produce bioelectricity and H2 gas at ambient temperatures. MFC- and MEC-type systems are close to the industrial application but are still in their infancy stage. General scale-up challenges for MFCs and MECs are finding low-cost electrode and membrane materials and implementing compact separated reactor designs (Montpart et al., 2015). Miller et al. (2019) reported that the steps for commercialization of MFCs and MEC technologies include starting from bench scales (experimental setup, reactors, electrode, and materials), designing, constructing and testing of the BES system, recording data and specifying important physicochemical parameters and thereafter optimizing the system (looking at system controls, recovery processes, and cost evaluation), and planning phases for demonstration at large scale. As an alternative to scale-up, an increase of productivities can be achieved by increasing the number of basic cell units (Ding et al., 2015), which is likely a better approach. Connecting them to stacked reactors in series or parallel can lead to improved power output and

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current generation (Ding et al., 2015). Nevertheless, the final cells in a series connection will have lower productivity rates due to a decrease in substrate concentrations (Heidrich et al., 2014; Miller et al., 2019). The performance of different reactor designs (e.g., tubular vs flat plate, separated vs nonseparated reactors) and different process modes (batch vs continuous) with varying volumes has been researched for wastewater treatment. Reengineering BES based on new insights and existing results into specific components (cell structure, electrodes, membranes, and hydraulic loads) can improve large-scale systems productivity. This can lead to broad applicable scaled-up BESs. To date, most of the systems investigated are tubular shaped and flat plate designs. As an alternative, systems with multiple electrodes can be used to ensure good positioning of components. There are also novel methods for avoiding the ohmic losses and edge effects that disturb the electric field and result in enhanced power densities (Heidrich et al., 2014; Miller et al., 2019). Stackable reactor systems seem to be the most suitable design concept, achieving target volumes by increasing the number of stacked units. These individual reactors can have well-characterized performances, which also allows for easy estimation of overall productivity. The most recent research projects and experiments have shown that the process performances at the technical scale are well below that of small-scale devices. Nevertheless, there is an industrial potential for MFCs and MECs. Reengineering based on existing results and new insights into the specific components would improve large-scale system productivity. These experiences of reengineering will certainly lead to some broad, applicable scale-up criteria for BES in general. In the case of hydrogen production, the price of H2 is more significant than the size of the MECs and electricity usage. Moreover, geometric surface area—electrode/reactor volume—ratio, and optimization of flow regime are important factors in designing MECs and MFCs (Khosravanipour Mostafazadeh et al., 2017). The environmental, economical, and technical feasibility of BES must be considered as a suitable long-term technology (Fig. 8.2). Reduced capital, maintenance, and operational costs need further analysis and compared to the conventional processes of wastewater treatment, and compatibility with environmental policies must be explored. Cusick et al. (2010) estimated that energy value recovery from hydrogen production from wastewater with MECs is US $0.19/kg COD and US $0.06/kg COD for domestic wastewater and winery wastewater, respectively. The electrode material and membrane are the most important factors for the techno-economic calculation of MECs and MFCs. Montpart et al. (2015) reported that membranes and cathode materials are solely responsible for up to 85% of the capital cost of BES (cathode: d500/m2 and membranes: d400/m2). Small- and medium-sized enterprises are interested in the capitalization of MECs and MFCs, and therefore cost-effective materials and design criteria are important factors for the feasibility of the commercialization of these systems (Fig. 8.2). The high capital costs of MECs and MFCs are currently the main obstructions to scale-up and industrialization. Moreover, to commercialize MEC and MFC processes, several parameters should be considered, such as product yield, final concentration of water, rates of

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Figure 8.2 Bioelectrochemical systems (MFCs or MECs integrated) with other processes as part of a bioeconomy substrate supply, product extraction and driven by renewable power (solar/wind/wave/geothermal energy). MECs, Microbial electrolysis cells; MFCs, microbial fuel cells. Source: Adapted after Jiang, Y., Jianxiong Zeng, R., 2018. Expanding the product spectrum of value added chemicals in microbial electrosynthesis through integrated process design—a review. Bioresour. Technol. 269, 503 512 (Jiang and Jianxiong Zeng, 2018).

product, and efficiency in terms of energy (Sharma et al., 2014; Kadier et al., 2016a,b; Khosravanipour Mostafazadeh et al., 2017). The final product concentration of stormwater or wastewater, as well as the substrate concentration, is also another crucial factor that determines the purification cost. The treatment rate relating to system volume also affects capital cost. Lastly, the energy efficiency (the amount of energy consumed per kg of product) needs to be reviewed to reduce the costs as well as high operating costs, such as the electricity and energy cost, substrate prices, and separation cost needs to be considered (Sharma et al., 2014; Kadier et al., 2016a; Khosravanipour Mostafazadeh et al., 2017).

8.5

Outlook, challenges, and future perspectives

BES can emerge as a potential alternative to combine stormwater and/or wastewater treatment with energy production (Fig. 8.3). Wastewater and BES technologies are currently symbiotically integrated to increase product yields and selectivity. Thereby, overall efficiencies to resolve the key issues with the upscaling of both MEC and MFC technologies are fundamental. They have shown great potential in producing value-added products inefficient ways with renewable sources under ambient conditions. Synergistic integrations with wastewater utilities and

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Figure 8.3 Sustainable water- and waste-derived bioeconomy prospects of the future.

stormwater management are imperative in attaining an economic and environmental upside of novel BES synthesis processes (Fig. 8.3). However, there are still several major challenges for the practical environmental engineering application. The production rate of MECs or MFCs could be improved by using more efficient microorganisms as well as novel electrode materials (Aryal et al., 2017; Chen et al., 2016; Tremblay and Zhang, 2015). There is an enormous gap in knowledge in the scaling-up of MFC and MEC reactors. As discussed previously, this scalingup of BES could potentially be achieved by stacking multiple reactors rather than simply enlarging a single reactor (Liang et al., 2018). The separation processes require the optimization of targeted pollutants. The separation of water contaminants in situ or via external recycle loops can be achieved based on various hydrophilic properties, as well as the difference of molecular electrical properties of ions in the system. Organic contaminants can be degraded, and nutrients or other ions recaptured. Some solutions to the problems and complications facing BES technologies can be enhanced by the integration of applied microbiology, engineering physics, and electrochemistry, alongside mathematical modeling and simulations presenting efficient systems (Aryal et al., 2017; Chen et al., 2016; Tremblay and Zhang, 2015). To optimize the energy generation and stormwater treatment, synergistic interaction between microbial strains and components should be studied and well understood (Kru¨ger et al., 2018). Assessments of the yields, physiochemical behavior, technical feasibility, and optimal conditions in achieving the highest cell efficiency (reduced overpotentials, mass-transfer losses) are required for further studies.

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There are economic constraints for electricity-driven bioproduction due to reactor construction, biological challenges, and the electricity needed, which could be considered for the pilot-plant design of BES (Chiranjeevi et al., 2018). Specific experimental designs in estimating efficiencies must take into account design parameters, including anode and cathode exchange and limiting current densities, anode, and cathode electrolytes, as well as the interconnectivity thicknesses and conductivity constants across the cells. Challenges of BES, such as low-voltage efficiencies, can be improved by decreasing overpotential and applied voltage to increase current densities across the cells (Givirovskiy et al., 2019). Furthermore, using suitable bacteria that are able to produce electrons from the substrate (such as enterobacter) can also be a viable alternative to enhance the efficacy in anode chambers, while reduction takes place at the cathode to produce solvents. Optimization microbial processes can have a positive effect on emerging biomass if implemented appropriately (Rago et al., 2019). Details about the electrical potential and current impact on the bacterial organism’s metabolism and microbial strains must be further understood. Real mixtures such as wastewater from industries or stormwater in laboratory scale and thereafter pilot-plant scale should be further explored. In addition, economic evaluations as part of a feasibility study on large-scale BES (MEC or MFC plants) should be carried out in greater detail. In addition, the cost of materials, efficient reactor configuration, and reaction rates are still the challenges for scaling-up BES (Chiranjeevi et al., 2018). Current and future research must be complemented by process integration research to facilitate the commercialization of BES technologies. More research is required on bioelectrochemical stormwater treatment, for example, on treatment efficiencies taking into account the varying contents and environmental conditions where stormwater is produced. Furthermore, the potential of simultaneously recovering nutrients or removing salts should be determined, and the microbial communities enriched with the anode-treating stormwater should be delineated. Low-cost and efficient reactor configurations and electrode and membrane materials that can be scaled up should be studied. One possibility for bioelectrochemical stormwater treatment could be bioelectrochemical-constructed wetlands that have been studied for the treatment of municipal wastewaters (Ramı´rez-Vargas et al., 2019). Current research must be complemented by process integration research to facilitate the commercialization of BES technologies.

8.6

Conclusion

Global environmental challenges, such as climate change, depletion of fossil fuel resources, and growing populations, demand concerted actions. Applying an integrated approach for the transition of oil-/gas-based economies toward a sustainable bioeconomy relying on biomasses-derived products is the future. The development of BESs such as MFCs or MECs is crucial for the implementation of a bioeconomy of the future. The technology-based innovations of BESs are at the center of the

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ongoing transformation processes. It appears to be essential to combine sustainable water, wastewater, and stormwater management practices with a circular bioeconomy approach to meet the social, economical, and environmental needs of the future. Capturing and using urban stormwater runoff for water harvesting can help alleviate water scarcity in semiarid regions. This is a new paradigm that views stormwater as a water source and not solely a flood or pollution problem. The approach to sustainable wastewater and stormwater treatment and management must be flexible enough to be adapted. Nutrients in stormwater and wastewater are renewable and alternative sources for traditional fertilizers. In addition, the organic compounds in storm- and wastewater can be transformed into electricity or energy carriers, such as hydrogen, while simultaneously producing cleaner water. In terms of maximizing productivity and minimizing the costs of large-scale BES, cheap, and readily available materials for electrodes and membranes are needed. In a technical electrochemical reactor, the use of expensive separators, such as membranes, should be avoided if possible. Bioelectrochemical reactors without a membrane, resulting in single-chamber architectures, simplify reactor design and can reduce capital costs. Novel systems are needed to reduce both the time required for experiments and the reactor costs, whilst allowing a high-throughput operation of the cells. BES still requires further improvements to get to larger scales and requires special reactor concepts tailored for their needs. Recent advances in new types of electrodes, a better understanding of the impact of components on performance, and results from several new pilot-scale tests are good indicators that broader commercialization of BES (MFC or MEC) technologies could become a reality within a few years’ time. While BES reactors for electrosynthesis need to be sterilizable, and the integration of downstream processing should be superior to conventional water and wastewater treatment processes, minimizing the costs of reactor components for bioelectrochemical cells needs to be addressed. For an industrial application of BES, both electron transfer principles should be kept in mind and considered from an economic and ecological point of view.

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Khan, M.Z., Nizami, A.S., Rehan, M., Ouda, O.K.M., Sultana, S., Ismail, I.M., et al., 2017. Microbial electrolysis cells for hydrogen production and urban wastewater treatment: a case study of Saudi Arabia. Appl. Energy 185, 410 420. Khosravanipour Mostafazadeh, A., Drogui, P., Brar, S.K., Tyagi, R.D., Bihan, Y.L., Buelna, G., 2017. Microbial electrosynthesis of solvents and alcoholic biofuels from nutrient waste: a review. J. Environ. Chem. Eng. 5 (1), 940 954. Ki, D., Popat, S.C., Torres, C.I., 2016. Reduced overpotentials in microbial electrolysis cells through improved design, operation, and electrochemical characterization. Chem. Eng. J. 287, 181 188 (70). Kiely, P.D., Cusick, R., Call, D.F., Selembo, P.A., Regan, J.M., Logan, B.E., 2011. Anode microbial communities produced by changing from microbial fuel cell to microbial electrolysis cell operation using two different wastewaters. Bioresour. Technol. 102 (1), 388 394. Kim, K., Yang, W., Evans, P.J., Logan, B.E., 2016. Continuous treatment of high strength wastewaters using air-cathode microbial fuel cells. Bioresour. Technol. 221, 96 101. Koo´k, L., Ro´zsenberszki, T., Nemesto´thy, N., Be´lafi-Bako´, K., Bakonyi, P., 2016. Bioelectrochemical treatment of municipal waste liquor in microbial fuel cells for energy valorization. J. Cleaner Prod. 112, 4406 4412. Krieg, T., Sydow, A., Schro¨der, U., Schrader, J., Holtmann, D., 2014. Reactor concepts for bioelectrochemical syntheses and energy conversion. Trends Biotechnol. 32 (12), 645 655. Krieg, T., Phan, L.M.P., Wood, J.A., Sydow, A., Vassilev, I., Kro¨mer, J.O., et al., 2018. Characterization of a membrane-separated and a membrane-less electrobioreactor for bioelectrochemical syntheses. Biotechnol. Bioeng. 115 (7), 1705 1716. Kru¨ger, A., Schaefers, C., Schroeder, C., Antranikian, G., 2018. Towards a sustainable biobased industry highlighting the impact of extremophiles. Nat. Biotechnol. 40 (Pt A), 144 153. Leong, J.X., Daud, W.R.W., Ghasemi, M., Liew, K.B., Ismail, M., 2013. Ion exchange membranes as separators in microbial fuel cells for bioenergy conversion: a comprehensive review. Renew. Sustain. Energy Rev. 28, 575 587. Liang, P., Duan, R., Jiang, Y., Zhang, X., Qiu, Y., Huang, X., 2018. One-year operation of 1000-L modularized microbial fuel cell for municipal wastewater treatment. Water Res. 141, 1 8. Liu, Y.P., Wang, Y.H., Wang, B.S., Chen, Q.Y., 2014. Effect of anolyte pH and cathode Pt loading on electricity and hydrogen co-production performance of the bioelectrochemical system. Int. J. Hydrogen Energy 39 (26), 14191 14195. Logan, B.E., 2010. Scaling up microbial fuel cells and other bioelectrochemical systems. Appl. Microbiol. Biotechnol. 85 (6), 1665 1671. Ohmic resistance in membranes. Logan, B.E., Hamelers, B., Rozendal, R., Schro¨der, U., Keller, J., Freguia, S., et al., 2006. Microbial fuel cells: methodology and technology. Environ. Sci. Technol. 40 (17), 5181 5192. Majumder, D., Maity, J.P., Tseng, M., Nimje, V.R., Chen, H., Chen, C., et al., 2014. Electricity generation and wastewater treatment of oil refinery in microbial fuel cells using Pseudomonas putida. Int. J. Mol. Sci. 15 (9), 16772 16786. Miller, A., Singh, L., Wang, L., Liu, H., 2019. Linking internal resistance with design and operation decisions in microbial electrolysis cells. Environ. Int. 126, 611 618. Mohan, S.V., Velvizhi, G., Krishna, K.V., Babu, M.L., 2014. Microbial catalyzed electrochemical systems: a bio-factory with multi-facet applications. Bioresour. Technol. 165, 355 364.

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Mohanakrishna, G., Vanbroekhoven, K., Pant, D., 2016. Imperative role of applied potential and inorganic carbon source on acetate production through microbial electrosynthesis. J. CO2 Util. 15, 57 64. Montpart, N., Rago, L., Baeza, J.A., Guisasola, A., 2015. Hydrogen production in single chamber microbial electrolysis cells with different complex substrates. Water Res. 68, 601 615. Nimje, V.R., Chen, C.Y., Chen, C.C., Chen, H.R., Tseng, M.J., Jean, J.S., et al., 2011. Glycerol degradation in single-chamber microbial fuel cells. Bioresour. Technol. 102 (3), 2629 2634. Pant, D., Singh, A., Van Bogaert, G., Gallego, Y.A., Diels, L., Vanbroekhoven, K., 2011. An introduction to the life cycle assessment (LCA) of bioelectrochemical systems (BES) for sustainable energy and product generation: relevance and key aspects. Renew. Sustain. Energy Rev. 15 (2), 1305 1313. Pant, D., Singh, A., Van Bogaert, G., Irving Olsen, S., Singh Nigam, P., Diels, L., et al., 2012. Bioelectrochemical systems (BES) for sustainable energy production and product recovery from organic wastes and industrial wastewaters. RSC Adv. 2 (4), 1248 1263. Pham, T.H., Rabaey, K., Aelterman, P., Clauwaert, P., De Schamphelaire, L., Boon, N., et al., 2006. Microbial fuel cells in relation to conventional anaerobic digestion technology. Eng. Life Sci. 6 (3), 285 292. Pisciotta, J.M., Zaybak, Z., Call, D.F., Nam, J.Y., Logan, B.E., 2012. Enrichment of microbial electrolysis cell biocathodes from sediment microbial fuel cell bioanodes. Appl. Environ. Microbiol. 78 (15), 5212 5219. Rabaey, K., Rozendal, R.A., 2010. Microbial electrosynthesis—revisiting the electrical route for microbial production. Nat. Rev. Microbiol. 8 (10), 706. Rago, L., Pant, D., Schievano, A., 2019. Electro-fermentation—microbial electrochemistry as new frontier in biomass refineries and industrial fermentations. Advanced Bioprocessing for Alternative Fuels, Bio-based Chemicals, and Bioproducts. Woodhead Publishing, pp. 265 287. Ramı´rez-Vargas, C.A., Arias, C.A., Carvalho, P., Zhang, L., Esteve-Nu´n˜ez, A., Brix, H., 2019. Electroactive biofilm-based constructed wetland (EABB-CW): A mesocosm-scale test of an innovative setup for wastewater treatment. Sci. Total Environ. 659, 796 806. Rosenbaum, M., Aulenta, F., Villano, M., Angenent, L.T., 2011. Cathodes as electron donors for microbial metabolism: which extracellular electron transfer mechanisms are involved? Bioresour. Technol. 102 (1), 324 333. Rozendal, R.A., Hamelers, H.V., Buisman, C.J., 2006. Effects of membrane cation transport on pH and microbial fuel cell performance. Environ. Sci. Technol. 40 (17), 5206 5211. Rozendal, R.A., Hamelers, H.V., Rabaey, K., Keller, J., Buisman, C.J., 2008a. Towards practical implementation of bioelectrochemical wastewater treatment. Trends Biotechnol. 26 (8), 450 459. Rozendal, R.A., Jeremiasse, A.W., Hamelers, H.V.M., Buisman, C.J.N., 2008b. Hydrogen production with a microbial biocathode. Environ. Sci. Technol. 42, 629 634. Sharma, M., Bajracharya, S., Gildemyn, S., Patil, S.A., Alvarez-Gallego, Y., Pant, D., et al., 2014. A critical revisit of the key parameters used to describe microbial electrochemical systems. Electrochim. Acta 140, 191 208. Shen, R., Jiang, Y., Ge, Z., Lu, J., Zhang, Y., Liu, Z., et al., 2018. Microbial electrolysis treatment of post-hydrothermal liquefaction wastewater with hydrogen generation. Appl. Energy 212, 509 515. Stager, J.L., Zhang, X., Logan, B.E., 2017. Addition of acetate improves stability of power generation using microbial fuel cells treating domestic wastewater. Bioelectrochemistry 118, 154 160.

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Thygesen, A., Thomsen, A.B., Possemiers, S., Verstraete, W., 2010. Integration of microbial electrolysis cells (MECs) in the biorefinery for production of ethanol, H 2 and phenolics. Waste Biomass Valoriz. 1 (1), 9 20. Membranes for isolation of gases in chamber. Tommasi, T., Ruggeri, B., Sanfilippo, S., 2012. Energy valorisation of residues of dark anaerobic production of hydrogen. J. Cleaner Prod. 34, 91 97. Tota-Maharaj, K., Paul, P., 2015. Performance of pilot-scale microbial fuel cells treating wastewater with associated bioenergy production in the Caribbean context. Int. J. Energy Environ. Eng. 6 (3), 213 220. Tremblay, P.L., Zhang, T., 2015. Electrifying microbes for the production of chemicals. Front. Microbiol. 6, 201. Trowsdale, S.A., Simcock, R., 2011. Urban stormwater treatment using bioretention. J. Hydrol. 397 (3), 167 174. Xie, B., Gong, W., Ding, A., Yu, H., Qu, F., Tang, X., et al., 2017. Microbial community composition and electricity generation in cattle manure slurry treatment using microbial fuel cells: effects of inoculum addition. Environ. Sci. Pollut. Res. 24 (29), 23226 23235. Yang, W., Kim, K.Y., Saikaly, P.E., Logan, B.E., 2017. The impact of new cathode materials relative to baseline performance of microbial fuel cells all with the same architecture and solution chemistry. Energy Environ. Sci. 10 (5), 1025 1033. Yu, F., Wang, C., Ma, J., 2016. Applications of graphene-modified electrodes in microbial fuel cells. Materials 9 (10), 807. Yu, Z., Leng, X., Zhao, S., Ji, J., Zhou, T., Khan, A., et al., 2018. A review on the applications of microbial electrolysis cells in anaerobic digestion. Bioresour. Technol. 255, 340 348. Zaidi, S.J., 2009. Research trends in polymer electrolyte membranes for PEMFC. Polymer Membranes for Fuel Cells. Springer, Boston, MA, pp. 7 25. Zhang, X., Cheng, S., Wang, X., Huang, X., Logan, B.E., 2009. Separator characteristics for increasing performance of microbial fuel cells. Environ. Sci. Technol. 43 (21), 8456 8461.

Further reading Guo, T., Sun, B., Jiang, M., Wu, H., Du, T., Tang, Y., et al., 2012. Enhancement of butanol production and reducing power using a two-stage controlled-pH strategy in batch culture of Clostridium acetobutylicum XY16. World J. Microbiol. Biotechnol. 28 (7), 2551 2558. Hiegemann, H., Herzer, D., Nettmann, E., Lu¨bken, M., Schulte, P., Schmelz, K., et al., 2016. An integrated 45L pilot microbial fuel cell system at a full-scale wastewater treatment plant. Bioresour. Technol. 218, 115 122. Khan, M.Z., Sim, Y.L., Lin, Y.J., Lai, K.M., 2013. Testing biological effects of handwashing grey water for reuse in irrigation on an urban farm: a case study. Environ. Technol. 34 (4), 545 551. Logan, B.E., 2009. Exoelectrogenic bacteria that power microbial fuel cells. Nat. Rev. Microbiol. 7 (5), 375. Pons, L., De´lia, M.L., Bergel, A., 2011. Effect of surface roughness, biofilm coverage and biofilm structure on the electrochemical efficiency of microbial cathodes. Bioresour. Technol. 102 (3), 2678 2683. Ucar, D., Zhang, Y., Angelidaki, I., 2017. An overview of electron acceptors in microbial fuel cells. Front. Microbiol. 8, 643.

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Abhilasha Singh Mathuriya1 and Soumya Pandit 2 1 Department of Life Sciences, School of Basic Sciences and Research, Sharda University, Greater Noida, India, 2Department of Life Sciences, Sharda University, Greater Noida, India

Chapter Outline 9.1 Introduction 199 9.2 Food-based wastes and wastewater as a substrate for microbial fuel cell 200 9.3 Beer brewery wastewater wastes and wastewater as a substrate for microbial fuel cell 202 9.4 Conclusion 203 References 206 Further reading 210

9.1

Introduction

Wastewater possesses an ample dose of complex biological and chemical materials, which may create serious health, sanitation, and environmental problems. These complex biological and chemical materials also allow the growth of enumerable microbial flora, which can survive in even extreme environment. Developing a track to harness electricity from the degradation of these biological and chemical materials by this microbial flora is the driving force for the development of microbial fuel cells (MFCs). MFCs are bioelectrochemical systems that directly convert chemical energy contained in organic matters present in wastewaters into electrical energy by utilizing the metabolic (catalytic) activity of microorganisms (Logan et al., 2006; Mathuriya and Sharma, 2009; Mathuriya and Pant, 2018). A prototype two-chamber MFC consists of one anaerobic anode chamber and one aerobic cathode chamber separated by a proton/cation-exchange membrane/separator (Mathuriya and Yakhmi, 2014; Huang et al., 2010; Mathuriya et al., 2018). At the anode, microorganisms liberate electrons by oxidation of the organic compounds. These liberated electrons move through the external circuit toward the cathode, while protons transferred through membrane or separator toward the cathode. At the cathode, these electrons and protons react with oxygen to form water (Allen and Bennetto, 1993). The electricity generated by this reaction can be used by an external resistor placed between anode and cathode (Jiang et al., 2011). These chambers can take various practical shapes for tailoring the architectural design, Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00009-6 © 2020 Elsevier Inc. All rights reserved.

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such as tubular, hexagonal, rectangular, serpentine, or micro-MFC (Scott et al., 2007; Zhuang et al., 2012; Mathuriya, 2016a,b,c,d). Single-compartment MFCs offer a more straightforward system. These are a single anodic chamber with no decisive cathode chamber. In such systems, cathodes are formed on the exterior side of the wall of the anodic chamber that utilizes oxygen from the atmosphere for oxygen reduction reaction using catalysts. These designs are easy to scale up and, therefore, extensively used for the study (Liu et al., 2004; Rozendal et al., 2006; Abourached et al., 2014; Mathuriya, 2016d; Logron˜o et al., 2017). One issue is the formation of aerobic biofilm on the cathode, which does not participate in the electroactivity, can be prevented by hydrophobic and oxygen permeable diffusion layer onto cathode (Yang et al., 2015), or by removing such biofilm (Rossi et al., 2018). Several design variations for MFCs have been reported to increase power density (PD) and for continuous flow through the anodic chamber (Tee et al., 2016; Chang et al., 2017; Taghavi et al., 2014; Sun et al., 2017; Mathuriya, 2016d). MFCs with membranes face some inherent limitations, such as the high cost of proton exchange membrane (PEM), biofouling, and high internal resistance (Liu and Logan, 2004; Liu et al., 2008). Membrane-less MFCs are one option, where often oxygen acts as the terminal electron acceptor. The elimination of membrane can simplify the MFC design and decreases the cost and the internal resistance. Several MFCs without membrane-like sediment MFCs (Tender et al., 2002; Thomas et al., 2013; Zhou et al., 2014), upflow membrane-less MFCs (Thung et al., 2015) have been studied in the past.

9.2

Food-based wastes and wastewater as a substrate for microbial fuel cell

Domestic, commercial, and industrial activities generate food waste, such as fruit and vegetable peels, unused food, and dish washings, which is usually disposed of in municipal waste. However, food waste contains rich organics fraction and high moisture, which makes it composite in nature, and food waste could be suitable as anolyte in MFCs. In one of the initial studies, Oh and Logan (2005) utilized cereal wastewater in a two-chambered MFC and gained up to 81 mW/m2 PD, per liter of wastewater with up to 95% chemical oxygen demand (COD) removal. While using one chambered MFC and prefermented wastewater, the maximum PD was 371 mW/m2. Patil et al. (2009) utilized chocolate industry wastewater as a substrate for electricity generation using activated sludge as inoculum in two-chambered MFC. The use of chocolate industry wastewater was shown to be promising with 4.1 mA current output along with a significant reduction in COD, biological oxygen demand (BOD), total solids, and total dissolved solids of wastewater by 75%, 65%, 68%, and 50%, respectively. You et al. (2010) studied the treatment of seafood wastewater in a continuous MFC with modified anoxic/oxic (A/O) architecture (A/O MFC) and achieved up to 16.2 W/m3 PD at a current density of 41.7 A/m3.

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Venkata Mohan et al. (2010a) treated composite waste vegetables extract in a single-chamber MFC with noncatalyzed electrodes. The system was operated with composite waste vegetables at three substrate load conditions (2.08, 1.39, and 0.70 kgCOD/m3 day), and the best power output of 57.38 mW/m2 was observed at lower substrate load. During operation, 62.86% COD, 79.84% carbohydrates, and 55.12% turbidity were also removed. Venkata Mohan and Chandrasekhar (2011) evaluated the feasibility of bioelectricity generation from composite canteen-based food waste in a solid-phase MFC. Distance between the electrodes and the presence of PEM marked significant influence on the power yields, and MFC in which anode placed 5.0 cm from cathode-PEM showed 463 mV, 170 mW/m2, and 76.0% COD reduction along with bioethanol production. Sangeetha and Muthukumar (2011) used food processing wastewater in a dual-chambered, salt-bridge MFCs with aerated catholyte and obtained a maximum PD of 123.8 mW/m2 and 98.9% COD removal. In an interesting study, Behera et al. (2010) used MFC fabricated using an earthen pot to treat rice mill wastewater at different feed pH. The best results were obtained at pH 8.0 with maximum COD removal efficiency of 96.5% and PD of 48.64 mW/m3 during 288 h of operation. Electricity generation from sweet potato-shochu waste was examined by Iigatani et al. (2019) in a cassette-electrode MFC (CE-MFC). Among CE-MFCs with raw (73 g-COD chromium CODCr/L) and different concentrations of diluted sweet potato-shochu waste (0.5, 1, 5, 10, and 20 g-CODCr/L) without pH control, the maximum PD (1.2 W/m3) and CODCr removal efficiency (67.4% 6 1.8%) were observed in the CE-MFCs with 10 g-CODCr/L shochu waste. The concentration of organic acid was decreased to below the quantification limits during the 9-day operation in the CE-MFC with 10 g-CODCr/L shochu waste. Cashew apple juice is an agro-based residue; it is a potential source of substrate for industrial and biotechnological processes to produce ethanol. Priya and Setty (2019) utilized clarified cashew apple juice as a substrate in a dual-chambered acrylic MFC and generated up to 0.4 V of open-circuit voltage and maximum PD and current density of 31.58 mW/m2, 350 mA/m2, respectively. In an interesting study, Oyiwona et al. (2018) utilized poultry droppings wastewater in MFC and achieved up to 6.9 6 3.1 W/m3 of volumetric PD. Noori et al. (2016) used air-cathode MFC having vanadium pentoxide microflowers catalyst on the cathode to treat fish market wastewater. MFCs demonstrated significant with 80% removal of COD and 60% protein removal efficiency along with PD of 6.06 W/m3. Jayashree et al. (2016) treated seafood processing wastewater in a tubular upflow MFC. At an organic loading rate (OLR) of 0.6 g/day, the MFC achieved up to 83% and 95% of total and soluble COD removal, respectively, while a maximum PD of 105 mW/m2 (2.21 W/m3) was achieved at an OLR of 2.57 g/day. In an interesting study, Subha et al. (2019) utilized the chocolate wastewater in upflow MFC to evaluate the effect of hydraulic retention time (HRT) and OLR on organic removal and power generation. Four bacterial species such as Achromobacter insuavis strain BT1 (MF346036.1), A. insuavis strain B3 (MF346037.1), Bacillus encimensis strain B4 (MF346038.1), and Kocuria flava strain B5 (MF346039.1) were shown to play a predominant role in substrate

202

Integrated Microbial Fuel Cells for Wastewater Treatment

degradation and power generation. A maximum PD of 98 mW/m2 and COD removal of 70% were obtained at an HRT of 15 h. The dairy industry is one of the prominent food-based industries of the world and involves the processing of raw milk into various edible products such as consumer milk, milk powder, condensed milk, cheese, butter, ice cream, yogurt, and many by-products, namely, buttermilk, whey, and their derivatives (Robinson, 1986). Hardly the dairy industry produces any solid waste. Since dairy wastewater contains biodegradable organics and nutrients (Venkata Mohan et al., 2007), it can be utilized as an efficient anolyte in MFCs. Venkata Mohan et al. (2010b) utilized dairy wastewater in a single-chamber noncatalyzed MFC and used anaerobic mixed consortia as anodic biocatalyst. MFC achieved 95.49% COD, 78.07% protein, 91.98% carbohydrates, and 99.02% turbidity along with 1.10 W/m3 PD, 308 mV voltage, and 1.78 mA current. Mathuriya and Sharma (2009) treated dairy wastewater in a two-chamber MFC and utilized the inherent microbial population of wastewater and achieved 10.89 mA current and 81.29% COD removal during 10 days of operation. Dalvi et al. (2011) utilized paneer whey in a two-chamber MFC using agar salt bridge and Klebsiella pneumoniae and obtained the maximum 453 mV open circuit voltage. In another interesting study, Nasirahmadi and Safekordi (2011) used Escherichia coli in a two-chamber MFC with humic acid as a mediator to generate electricity from cheese whey and achieved a maximum 324.8 μW power and 1194.6 μA current during operation. Mansoorian et al. (2016) developed a catalyst-less and mediator-less membrane MFC and observed dairy industry wastewater treatment in it. The maximum current intensity and PD produced were, respectively, 3.74 mA and 621.13 mW/m2 on the anode surface, at OLR equal to 53.22 kgCOD/m3 day and at the external resistance of 1 kΩ. The maximum voltage produced was 0.856 V at OLR equal to 53.22 kgCOD/m3 day and at temperature 35 C. The maximum coulombic efficiency of 37.16% was achieved at OLR equal to 17.74 kgCOD/m3 day. In an interesting study, Faria et al. (2017) developed a continuous MFC to treat dairy industry wastewater. During 20 days of operation, a maximum voltage of 576 mV was produced while PD reached up to 1.9 W/m3. MFC also removed 63% 6 5% COD at a hydraulic retention time of 8.4 h. Sekar et al. (2019) evaluated the performance of MFC in treating carbohydrates and proteins containing dairy effluent and achieved a peak PD of 161.5 mW/m2 and COD removal efficiency up to 75%, respectively.

9.3

Beer brewery wastewater wastes and wastewater as a substrate for microbial fuel cell

Beer breweries generate wastewater from cooling (e.g., saccharification cooling, and fermentation) and washing units, causing several environmental problems. Brewery wastewater has high COD value but is a suitable substrate for microbes as much of the organic matter in the wastewater consists of sugar, starch, and protein (Gil et al., 2003). Many investigators studied the treatment of beer brewery

Treatment of food processing and beverage industry wastewaters in microbial fuel cells

203

wastewater in MFCs. Feng et al. (2008) achieved up to 528 mW/m2 PD and 98% COD removal from beer brewery wastewater an air-cathode MFC when 50 mM phosphate buffer was added to the wastewater. Wang et al. (2008) reported 87% COD removal and 483 mW/m2 PD from beer brewery wastewater in a singlechamber, membrane-free MFC. Wen et al. (2010) achieved up to 91.7% 95.7% of COD removal efficiency using brewery wastewater in a sequential anode cathode double-chamber MFC along with 830 mW/m3 PD. Mathuriya and Sharma (2010a, b) achieved 10.89 mA current and 93.8% COD removal efficiency by treatment of brewery wastewater in a two-chambered MFC. Anupama et al. (2011) used doublechambered MFC for the treatment of distillery wastewater, and a maximum COD removal of 64% along with a maximum PD of 18.35 mW/m2 was achieved during operation. Angosto et al. (2015) conducted an experiment in a single-cell air-cathode MFC in batch and fed-batch modes. They observed that brewery wastewater (CV1) mixed with pig-farm liquid manure gave the highest voltage (199.8 mV) and PD (340 mW/m3) outputs than nonmixed brewery wastewater with 53% COD removal. However, COD removal was higher (93%) when a pure brewery sample was used. Penteado et al. (2017) evaluated the treatment of winery wastewater in MFCs with different electrode materials. MFC with carbon felt reached the highest voltage and power 72 mV and 420 mW/m2, respectively, along with organic matter consumption rate of 650 mg COD/L/day. Lu et al. (2017) constructed a 20-L continuous-flow MFC containing two 10 L MFC reactors for high COD brewery wastewater (3196 6 978 mg/L) treatment over a year. These MFCs were able to treat up to 94.6% 6 1.0% COD during the first 150 days and CE . 5.5% during the first 286 days, along with the highest PD of 1.61 mW/m2 during operation.

9.4

Conclusion

MFC technology has emerged as a competent technology for simultaneous wastewater treatment and electricity generation. Recently, MFCs gained the attention of researchers, and the performance of MFCs has increased remarkably during the last two decades. This technology is spreading its aura in many directions, including wastewater treatment, sensor applications, heavy metal recovery, nitrification and denitrification, robotics, in situ power supply, and implantable power sources. Among all, wastewater treatment capabilities of MFCs are most popular. MFCs hold promise toward sustainable power generation and food-based wastewater treatment. Researchers across the globe are working on MFCs almost on every aspect of its performance. Many companies have started developing MFC reactors and playing well in the market. Some companies (MFC tech, Opencel Canada, Emefcy, Israel) have emerged to use MFC technology, and this technology could have a greater impact on the development of clean energy in the coming years (Table 9.1).

Table 9.1 Waste treatment efficiency of microbial fuel cells (MFCs) with a wide range of wastes. S. N.

Wastewater/ contaminant type

MFC architecture

COD/waste removal

Bioelectricity produced

References

1

Cassava

Two chamber

88%

1771 mW/m2

2 3

Two chamber Single chamber

75% 62.86%

3.02 A/m2 57.38 mW/m2

Solid phase

76%

170.81 mW/m2

Single chamber

64.83%

390 mA/m2

6

Chocolate industry Composite vegetable waste Composite canteen-based food waste Canteen-based composite food waste Seafood

Kaewkannetra et al. (2011) Patil et al. (2009) Venkata Mohan et al. (2010a) Venkata Mohan and Chandrasekhar (2011) Goud et al. (2011)

16.2 W/m3

You et al. (2010)

7 8

Cereal Meatpacking

95% 86% BOD

371 mW/m2 80 mW/m2

9 10 11

Food processing washdown water Dairy industry Dairy industry

84% of the soluble COD 63 6 5% 53.22 kgCOD/m3 day

Oh and Logan (2005) Heilmann and Logan (2006) Boghani et al. (2017)

1.9 W/m3 621.13 mW/m2

12 13 14 15

Raw dairy manure Chocolate Seafood processing Food processing

70% 95% COD 98.9%

138 6 19 mW/m2 98 mW/m2 105 mW/m2 123.8 mW/m2

Faria et al. (2017) Mansoorian et al. (2016) Powers et al. (2011) Subha et al. (2019) Jayashree et al. (2016) Sangeetha and Muthukumar (2011)

4

5

Continuous MFC with modified A/O architecture Single chamber Single chamber Tubular Continuous Catalyst-less and mediator-less membrane Column Upflow Tubular upflow Two chamber, salt bridge

16

Whey

Two chamber

17

Whey

Two chamber

18

Dairy

Sediment

19

Dairy

20

Molasses

21

390 6 21 W/m2

44.6% in the second cycle 92.8%

1800 6 120 W/m2

Single chamber

72.85%

52.92 3 1026 mW/ cm2 251 mA/m2

Molasses

Upflow anaerobic sludge blanket reactor-MFC-biological aerated filter Anaerobic baffled stacking

53.2% COD, 52.7% sulfate, 41.1% color 50% 70%

22 23 24 25 26 27

Beer brewery Brewery Beer brewery Brewery Alcohol distillery Distillery

Single chamber, membrane free Single chamber Single chamber Sequential anode cathode TC-MFC Combined AFB and MFC Single chamber

28 29 30

Distillery Brewery Brewery

Two chamber Continuous flow Single chamber

87% 43% 98% 92.2% 95.1% 80% 90% 72.84% COD, 31.67% color, 23.96% TDS 64% 94.6 6 1.0% 53% COD

1410.2 mW/m2

115.5 6 2.7 mW/ m2 483 mW/m2 264 mW/m2 528 mW/m2 830 mW/m3 124.03 mW/m2 124.35 mW/m2 18.35 mW/m2 1.61 mW/m2 340 mW/m3

A/O, Anoxic/oxic; AFB, anaerobic fluidized bed; COD, chemical oxygen demand; TC-MFC 5 two-chamber MFC; TDS, total dissolved solids.

Kassongo and Togo (2011a) Kassongo and Togo (2011b) Saravanan et al. (2010) Velasquez-Orta et al. (2011) Zhang et al. (2009)

Zhong et al. (2011) Wang et al. (2008) Wen et al. (2009) Feng et al. (2008) Wen et al. (2010) Huang et al. (2011) Mohanakrishna et al. (2010) Anupama et al. (2011) Lu et al. (2017) Angosto et al. (2015)

206

Integrated Microbial Fuel Cells for Wastewater Treatment

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Wen, Q., Wu, Y., Cao, D., Zhao, L., Sun, Q., 2009. Electricity generation and modeling of microbial fuel cell from continuous beer brewery wastewater. Biores. Technol. 100, 4171 4175. Wen, Q., Wu, Y., Zhao, L., Sun, Q., Kon, F., 2010. Electricity generation and brewery wastewater treatment from sequential anode-cathode microbial fuel cell. J. Zhejiang Univ. Sci., B 11, 87 93. Yang, W., Kim, K.Y., Logan, B.E., 2015. Development of carbon free diffusion layer for activated carbon air cathode of microbial fuel cells. Biores. Technol. 197, 318 322. You, S.J., Zhang, J.N., Yuan, Y.X., Ren, N.Q., Wang, X.H., 2010. Development of microbial fuel cell with anoxic/oxic design for treatment of saline seafood wastewater and biological electricity generation. J. Chem. Technol. Biotechnol. 85, 1077 1083. Zhang, B., Zhao, H., Zhou, S., Shi, C., Wang, C., Ni, J., 2009. A novel UASB-MFC-BAF integrated system for high strength molasses wastewater treatment and bioelectricity generation. Biores. Technol. 100, 5687 5693. Zhong, C., Zhang, B., Kong, L., Xue, A., Ni, J., 2011. Electricity generation from molasses wastewater by an anaerobic baffled stacking microbial fuel cell. J. Chem. Technol. Biotechnol. 86, 406 413. Zhou, Y.L., Yang, Y., Chen, M., Zhao, Z.W., Jiang, H.L., 2014. To improve the performance of sediment microbial fuel cell through amending colloidal iron oxyhydroxide into freshwater sediments. Biores. Technol. 159, 232 239. Zhuang, L., Yuan, Y., Wang, Y., Zhou, S., 2012. Long-term evaluation of a 10-liter serpentine-type microbial fuel cell stack treating brewery wastewater. Biores. Technol. 123, 406 412.

Further reading Logan, B.E., Murano, C., Scott, K., Gray, N.D., Head, I.M., 2005. Electricity generation from cysteine in a microbial fuel cell. Water Res. 39, 942 952. Mathuriya AS. 2017. A Microbial Fuel Cell With Interconnected Anode Chambers. Indian Patent: 201711046878:27.12.2017. Shukla, A.K., Suresh, P., Berchmans, S., Rajendran, A., 2005. Biological fuel cells and their applications. Curr. Sci. 87, 455 468.

Microbial fuel cell coupled with microalgae cultivation for wastewater treatment and energy recovery

10

P. Chiranjeevi and Sunil A. Patil Department of Earth and Environmental Sciences, Indian Institute of Science Education and Research Mohali (IISER Mohali), Sector 81, S.A.S. Nagar, Manauli, Punjab, India

Chapter Outline 10.1 Introduction

213

10.1.1 Microbial fuel cells 213 10.1.2 Microalgae cultivation 214

10.2 Microbial fuel cell and microalgae cultivation
based integrated systems 215 10.2.1 Microbial fuel cells coupled with the algal photobioreactors 215 10.2.2 Microbial fuel cells with the algal biocathodes 218

10.3 Factors influencing the performance of integrated microbial fuel cell and microalgae cultivation systems 221 10.3.1 10.3.2 10.3.3 10.3.4

Light intensity 221 Carbon dioxide 222 pH 222 Dissolved oxygen 223

10.4 Conclusion 223 Acknowledgment 223 References 224

10.1

Introduction

10.1.1 Microbial fuel cells Microbial fuel cell (MFC) technology, which is one of the prominent microbial electrochemical technologies, is considered as a promising approach for wastewater treatment and power generation (Logan and Regan, 2006; Venkata Mohan et al., 2008). The possibility of removing organic carbon, low levels of energy production and less sludge production makes the MFC technology promising for the treatment of various types of wastewaters (Holzman, 2005; Xie et al., 2011; Pandey et al., 2016). However, low nitrogen and phosphorus removal efficiency, electrochemical losses, costly reactor components, membrane fouling, and terminal electron Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00010-2 © 2020 Elsevier Inc. All rights reserved.

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acceptor (TEA) availability at the cathode are some of the limiting factors of this technology (Santoro et al., 2017).

10.1.2 Microalgae cultivation Photosynthetic microalgal cultivation utilizes the nutrients such as nitrogen and phosphorus along with organic or inorganic carbon present in the wastewaters and solar energy to produce useful biomass (High, 1996; Venkata Mohan et al., 2014a,b). Microalgae cultivation in oxidation ponds, raceway ponds, and high rate algal bioreactors has been shown to remove high concentration of nutrients, heavy metals, and even toxins from the wastewaters (Munoz and Guieysse, 2006; Richardson et al., 2012; Murphy et al., 2013). Microalgae can assimilate additional amount of phosphates and store in the form of polyphosphate granules thereby contributing to a high removal of phosphates from the wastewaters (Wang et al., 2010). They uptake the nitrogen majorly in the form of ammonia (NH1 4 ) and nitrates 2 (NO2 ) followed by nitrite (NO ) and urea (Larsdotter et al., 2006a). Chlorella vul3 2 garis, Chlorella pyrenoidosa, Spirulina platensis, Chlorella zofingiensis, Scenedesmus spp, and Chlorococcum spp. RAP13 have been reported to treat various complex wastewaters, namely, poultry litter leachates, olive-oil mill wastewater, and dairy wastewater and achieve decent removal of carbon (in the form of chemical oxygen demand (COD) removal up to 73%), phosphates (95%) and nitrogen (75%
85%) with significant biomass (5.8 g/L) and lipid (39%
42%) productivity (Beevi and Sukumaran, 2014; Li et al., 2012; Markou et al., 2012a,b, 2016). Other microalgal species such as Scenedesmus obliquus (Ji et al., 2013), Scenedesmus acutus (Sacrista´n de Alva et al., 2013), Botryococcus braunii (Orpez et al., 2009), and Auxenochlorella protothecoides (Hu et al., 2012) have also been reported to remove 70%
95% of phosphates and nitrates with lipid accumulation of 17%
20%. Microalgae assimilate acetate through its metabolic process [acetylcoenzyme A (acetyl-CoA) catalyzed by acetyl-CoA synthetase] and convert into high carbon lipids (Boyle and Morgan, 2009; Devi et al., 2012). The use of algae-based approaches for wastewater treatment makes the whole process sustainable and economic (Stephens et al., 2010; Venkata Mohan et al., 2014a,b). Due to its ability to utilize and thus remove nutrients as well as organic acids from effluents, microalgae cultivation approach has been tested along with other wastewater treatment approaches. For instance, microalgae cultivated in both mixotrophic and heterotrophic nutrition modes on carboxylic acid rich effluents from acidogenic hydrogen producing bioreactors resulted in high biomass and lipid productivities (Chiranjeevi and Venkata Mohan, 2017). Its integration with MFCs has also been tested for improving the overall wastewater treatment process (Kruzic and Kreissl, 2009). This chapter focuses on the integration of MFC technology with microalgae cultivation for improving the wastewater treatment and energy recovery. It does not cover other photosynthetic approaches, namely, bacterial photo-bioanodes and photo-biocathodes, and plant-based MFCs (Lyautey et al., 2011; Cai et al., 2013; Strik et al., 2008; Chiranjeevi et al., 2012).

Microbial fuel cell coupled with microalgae cultivation for wastewater treatment

10.2

215

Microbial fuel cell and microalgae cultivation
based integrated systems

The microalgal photosynthetic machinery is capable of converting solar energy to chemical energy by utilizing inorganic or organic carbon, nutrients, and water, and releasing oxygen into the atmosphere. Based on these metabolic capabilities, two different approaches have been reported for integrating MFC and microalgae cultivation processes (Table 10.1). The first approach is based on feeding the partially treated effluent from the anode chamber of MFCs to the microalgae cultivation reactor. The goal, in this case, is to achieve complete utilization of the nutrients and thereby to enhance the overall wastewater treatment. The second approach is based on the use of algal biocatalysts at the cathode of MFCs for creating nonlimiting TEA conditions through the continuous release of oxygen. In this case the cathode is referred to as the algal biocathode.

10.2.1 Microbial fuel cells coupled with the algal photobioreactors In this approach, both MFC and microalgal photobioreactor systems are operated independently of each other. Jiang et al. reported an integrated system by coupling upflow membrane-less MFC with a photobioreactor (Fig. 10.1A) (Jiang et al., 2013). The domestic wastewater fed to the MFC reactor was allowed to flow from the anode to the cathode chamber and eventually to the connected photobioreactor. It resulted in enhanced wastewater treatment (77.9% COD removal) efficiency and power production (power density, 481 mW/m3). Importantly, the nutrients containing MFC effluent fed to the algal cultivation system showed more than 90% removal of total phosphates and ammoniacal nitrogen (NH1 4 -N). This study thus demonstrated the feasibility of coupling MFCs with the microalgal cultivation system for efficient domestic wastewater treatment with concurrent power and biomass production. Gajda et al. (2013) integrated algal-biocathode chamber of MFC to microalgal cultivation system and recirculated algal biomass to the cathode for active oxygenation (Fig. 10.1B). Continuous recirculation of algal biomass enhanced oxygen reduction reaction (ORR) at the cathode which resulted in high power yields. Another study reported on coupling algal bioreactor to the open-air cathode MFC for high oxygen supply to the cathode (Kakarla et al., 2015) (Fig. 10.1C). Oxygen evolved from the algal reactor was nearly 30% higher than the atmospheric oxygen concentration in this case. It resulted in the efficient ORR at the cathode thereby contributing to a high power density. Tse et al. (2016) reported on the coupling of MFC with the membrane photobioreactor. This integrated approach resulted in maximum COD (97%) and ammonia (100%) removal efficiency and also showed increment in biomass production (up to 133 mg/L). The energy analysis also revealed that net energy production was high with the algal biomass (0.033 kWh/m3) than direct energy production through MFC showing positive energy balance of the systems. Jiang et al. fed the microalgal cultivation system

Table 10.1 Performance of microbial fuel cell (MFC) and microalgae cultivation
based integrated systems. S. no.

Type of integration

Biocathode

Bioanode

COD , P and N removal (%)

Power outputa

Reference

1

MFC coupled with algal photobioreactor

Mixed microalgae

Activated sludge

481 mW/m3

Jiang et al. (2013)

Chlorella vulgaris Mixed microalgae Mixed microalgae

Enriched mixed culture Anaerobic sludge Anaerobic sludge

630 mW/m2 3000 mW/m3 268 mW/m2

Gajda et al. (2013) Tse et al. (2016) Jiang (2017)

MFC with algal biocathodes

Pseudokirchneriella subcapitata Chlorella pyrenoidosa

Anaerobic sludge

COD, 77.9; P, 99.3; N, 99.0
COD, 97; N, 100 COD, 67; P 97; N, 99 COD, 92; N, 98; P, 82 COD, 87.3

2200 mW/m3

Xiao and He (2014)

6400 mW/m3

Jadhav and Jain (2017) Juang et al. (2012)

2 3 4 5 6 7 8 9

Chlorella spp. and Phormidium spp. Mixed microalgae Mixed microalgae

10 11 12

C. vulgaris Leptolyngbya sp. Scenedesmus obliquus

13

Mixed microalgae

14 15

C. vulgaris C. vulgaris

a

Normalized to volume of the reactor (m3) or to the anode surface area (m2). COD, Chemical oxygen demand.



Mixed anaerobic sewage sludge Mixed culture aerobic sludge Anaerobic sludge Anaerobic sludge Enriched mixed culture Leptolyngbya sp. Anaerobic bacteria from wastewater Anaerobic bacteria from wastewater Anaerobic sludge Pretreated cow manure

COD, 80

1.8 mW/m2

COD, 68.3 COD, 74.2

12.6 mW/m2 57 mW/m2

COD, 80 COD, 43.9


24.4 mW/m2 0.0008 mW/m2 30 mW/m2

COD, 97; N, 98

0.011 V MFC voltage 19151 mW/m3 67.07 mW/m2

COD, 44 COD, 70.8

Lobato et al. (2013) Venkata Mohan et al. (2014a,b) Wu et al. (2013) Maity et al. (2014) Kakarla and Min (2014) Nguyen et al. (2017) Hou et al. (2016) Khandelwal et al. (2018)

Microbial fuel cell coupled with microalgae cultivation for wastewater treatment

217

(A)

Treated wastewater + CO2

Cathode Glass wool

Anode

Photobioreactor

Resistor CO2

Wastewater

MFC

(B) Resistor

Wastewater

Cathode

Anode MFC

Photobioreactor

Figure 10.1 MFC and microalgae cultivation
based integrated systems: (A) upflow mode, membrane-less MFC cathode linked to algal photobioreactor, (B) MFC with the algal biocathode, (C) open-air cathode MFC linked to algal photobioreactor, and (D) MFC anode linked to algal photobioreactor. MFC, Microbial fuel cell.

with partially treated wastewater (COD, 67%; total phosphates, 34%; and NH1 4 -N, 50%) from the single-chambered MFC and reported a high removal of total phosphates (97%) and NH1 4 -N (99%) (Fig. 10.1D) (Jiang, 2017). The MFC system showed the highest power density of 268.5 mW/m2 because of the coupled algal cultivation system. All these studies demonstrated that the strategy of integrating both the systems could be an effective approach for renewable energy generation and energy-efficient wastewater treatment in a closed-loop biorefinery model (Tse et al., 2016).

218

Integrated Microbial Fuel Cells for Wastewater Treatment

(C)

Cathode

Resistor

Wastewater

O2

Anode Duelchambered MFC

Photobioreactor

(D)

Resistor

Wastewater

Cathode

Anode

Photobioreactor

MFC Effluent

Figure 10.1 (Continued)

10.2.2 Microbial fuel cells with the algal biocathodes Microalgae metabolism is flexible in terms of growth under various nutritional modes such as autotrophic, mixotrophic, and heterotrophic conditions depending on the availability of light and carbon source (Cuaresma et al., 2009; Venkata Mohan et al., 2014a,b). The mixotrophic mode of cultivating microalgae in the cathodic chamber of MFC offers operational flexibility due to logarithmic growth rate, efficient organic and inorganic (CO2) carbon assimilation, and capability to produce oxygen by algae (Cuaresma et al., 2009). In MFCs with the algal biocathodes, O2 produced through oxygenic photosynthesis acts as a TEA. Microalgae also utilize

Microbial fuel cell coupled with microalgae cultivation for wastewater treatment

219

CO2 and produce biomass (El-Mekawy et al., 2014; Campos-Martin et al., 2006; Powell et al., 2009). The oxygenic photosynthetic reactions of algal biocathode are as shown next. 6CO2 1 12H1 1 12e2 ! C6 H12 O6 ðbiomassÞ 1 3O2 ðlight reactionÞ O2 1 4H1 1 4e2 ! 2H2 Oðcathodic reactionÞ The O2 generated through self-regenerative microalgal biomass growth in the cathodic chamber is a potential alternative to mechanical aeration in MFCs. The use of algal biocathodes helps in enhancing the electrogenesis, CO2 sequestration (in the form of microalgal products) and wastewater treatment (through anodic substrate oxidation and cathodic reduction reactions) (Venkata Mohan et al., 2014a,b). Several studies have reported the advantages of the use of the algal-biocathode approach in MFCs. For instance, Xiao and He (2014) developed an integrated system by inserting MFC in the microalgal photobioreactor and operated it with synthetic wastewater for 1 year and evaluated treatment efficiency with simultaneous bioenergy production (Fig. 10.2A). The system showed active organic carbon (92%, COD) and nutrients removal (98% of NH1 4 -N and 82% of phosphates) along with the production of the power density of 2.2 W/m3 and 128 mg/L algal biomass. Jadhav and Jain (2017) also operated the similar type of setup using two different algal species C. pyrenoidosa and Anabaena ambigua along with the control (without algae) and evaluated wastewater treatment and electricity production. MFC system submerged in the C. pyrenoidosa culture showed high power density (6.4 W/m3) and biomass production (0.066 g/L/day) with COD removal of 87.3%. Cathodic biofilm formed by C. pyrenoidosa resulted in high oxygen evolution in the cathode chamber showing decrement in the internal resistance of the system due to efficient reduction of TEA than A. ambigua (low oxygenic algae) biocathode. Juang et al. (2012) compared the effects of different light power illumination on microalgal growth, electricity generation, and wastewater treatment by feeding with different concentrations of synthetic wastewater (982 and 1266 mg/L) in the cathode chamber and resulted maximum power density (1.8 mW/m2) and 80% COD removal. Another reactor configuration based on the direct inoculation of algae in the cathode chamber of MFC for continuous oxygen production has also been reported (Fig. 10.2B) (Lobato et al., 2013; Venkata Mohan et al., 2014a,b). These studies showed sustainable power generation (power density, 12.6 and 57 mW/m2) and wastewater treatment (COD, 68.3%, and 74.2%) by avoiding mechanical aerator and use of external mediators making the overall process more economical. Several researchers eventually tested various pure algal cultures in the cathode chamber of MFCs. For instance, Wu et al. (2013) used C. vulgaris in the modified MFC cathode in a tubular photobioreactor and evaluated the system performance with two different cathode materials during light and dark variation cycles. It achieved a power density of 24.4 mW/m2 with COD removal efficiency

220

Integrated Microbial Fuel Cells for Wastewater Treatment

(A)

Resistor ee-

e-

CO2 H+

H+

e-

CO2

Anode

Cathode

eAlgal cultivation system

(B)

Resistor

Wastewater

Anode

Cathode

Dual-chambered MFC

Figure 10.2 MFC with algal-biocathode system designs: (A) submerged MFC in photobioreactor (B) dual-chamber algal-biocathode MFC. MFC, Microbial fuel cell.

of nearly 80%. Prakash et al. (2014) used Leptolyngbya sp. at the cathode and reported 3.3 g/L biomass production (produced at a rate of 0.47 g/L/day), 1.06 g/g biomass of lipid production (containing carbon number C16:0, C18:2n-6, C18:1, and C16:1), power density of 0.008 mW/cm2, and overall COD removal of 43.9% within 7 days of operation. Kakarla and Min (2014) evaluated the different cathode electrode materials (carbon fiber brush and plain carbon paper) and concluded that carbon fiber brush at cathode resulted in high voltage generation (0.21 6 0.01 V) with a maximum power density of 30 mW/m2. In another study by the same group, treatment of landfill leachate using algal-biocathode MFC showed efficient COD

Microbial fuel cell coupled with microalgae cultivation for wastewater treatment

221

removal (97%) along with considerable voltage output (0.3 6 0.011 V) (Nguyen et al., 2017). Hou et al. (2016) operated algal-biocathode MFC using C. vulgaris for the treatment of organic-rich food waste by varying the inoculum densities. With an initial inoculum size of 0.15 g/L, the system resulted in the effective COD removal (44%), power output (19,151 mW/m3), and biomass productivity (0.2 g/L/day) with additional lipids production. Recently, Khandelwal et al. (2018) studied algal-biocathode MFC in a closed-loop approach. The algal biomass grown in the cathode chamber was harvested and extracted for lipids, and deoiled biomass was then used as a substrate in the anode chamber of MFC. The overall energy harvested in the form of electricity and bio-oil was around 0.0136 kWh kg/COD/day and 0.0782 kWh/m3/day, respectively. All these studies successfully demonstrated the sustainable way of using algal biocathodes instead of artificial aeration for the cathodic ORR in MFCs (Powell et al., 2009; Gajda et al., 2013).

10.3

Factors influencing the performance of integrated microbial fuel cell and microalgae cultivation systems

Apart from the reactor configurations and integration strategies, operational conditions, such as light intensity, carbon source, pH, and dissolved oxygen (DO), can influence the algal biomass growth and thus the overall performance of the integrated system.

10.3.1 Light intensity Illuminating subsaturating light intensity results in the enhanced photosynthetic reaction center proteins and chlorophyll pigments (Larsdotter, 2006a,b; Shivkumar et al., 2012). However, photooxidation in algal cells can also occur at oversaturated light intensities leading to the denaturation of photosynthetic pigments and proteins (Larsdotter, 2006a,b; Shivkumar et al., 2012). Campo et al. (2013) investigated the performance of algal biocathode by differing light exposure times (12 h light and 12 h dark cycles). Exposure for 12 h light showed efficient oxygen production due to active oxygenic photosynthesis of microalgal biomass. During the dark period, microalgae utilize oxygen produced in oxygenic photosynthesis and oxidize the organic matter. In this study, enhanced power generation and additional carbon removal during the dark cycle were reported. Two different views have emerged concerning the use of light intensity in the algal-biocathode systems. Wu et al. (2014) suggested that a high light intensity (3500 lx) with optimum light and dark cycle increases power output and wastewater treatment. Whereas Juang et al. (2012) recommended that a low light power (968
1937 lx) illumination is best suitable for enhanced power generation with considerable organic removal (4.41 mW/kg COD/day). Wang et al. (2010) observed that illumination at a

222

Integrated Microbial Fuel Cells for Wastewater Treatment

high light intensity reduces the power production of algal-biocathode MFC. They suggested the chances of oxygen crossover from the cathode to the anode chamber due to oxygen gradient thereby affecting the bioanode performance and thus overall performance of the system. Illuminating a very high light intensity to the algal biocathode can also bleach out the photosynthetic pigments and proteins leading to photo-inhibition (Wang et al., 2010).

10.3.2 Carbon dioxide Microalgae are capable of direct utilization of organic carbon (e.g., present in wastewaters) by heterotrophic and mixotrophic growth modes and inorganic carbon (CO2) through the phototrophic and mixotrophic growth modes. Atmospheric CO2 dissolved in the water and converted into bicarbonate/carbonate is one of the most abundant forms of buffer that exists in the natural waters. The inorganic form of carbon either as CO2 or HCO2 3 is interconvertible in the presence of carbonic anhydrase enzyme present in all microalgal species (Larsdotter, 2006a,b). Studies on CO2 diversion from the anode to the cathode chamber suggest that continue CO2 sparging is not required for active algal growth to produce sufficient oxygen (Cao et al., 2009; Wang et al., 2010; Cui et al., 2014). Even sparging of pure CO2 for 30 min into the cathode chamber also showed optimum growth of algal biomass (Campo et al., 2013). In the presence of light, microalgae preferably choose CO2 rather than organic carbon contributing to high power output during the day times (Myerse, 1980). Increasing the initial inoculum size of microalgae in the system overcomes the limitation associated with a lowered pH of the water due to CO2 solubility forming H2CO3 (Zhang et al., 2014; Chiu et al., 2009). Increase in the CO2 concentration impacts biomass productivity and lipid production (Liu et al., 2011). For instance, an increased supply of CO2 (10%
15%) showed 6% increment in the lipid production (Ho et al., 2010), but it did not impact the quality of the lipids (Mehrabadi et al., 2016).

10.3.3 pH The metabolic and biochemical processes of microalgae are also affected by the pH of growth medium or wastewater. Algal biomass grows effectively at the pH of 8 (Venkata Mohan et al., 2014a,b). The bicarbonate formation due to the dissolution of CO2 will raise the pH above 10 in most of the natural aquatic environments. An increase of pH to even 11 has been observed when inorganic carbonates were used as the carbon source (Devi and Venkata Mohan, 2012). The pH of the medium or wastewater also affects the uptake of nitrogen during microalgal growth. The pH can drop to 3 because of the ammonia assimilation by microalgae. Whereas utilization of nitrates raises the pH of the medium to alkalinity. Therefore the pH of the medium controls not only the growth of microalgae but also the nitrogen removal efficiency in the wastewaters (Park et al., 2011). Thus controlling pH is essential to the microalgae cultivation processes.

Microbial fuel cell coupled with microalgae cultivation for wastewater treatment

223

10.3.4 Dissolved oxygen DO concentration in algal-biocathode chambers impacts the overall MFC performance (Biffinger et al., 2008). DO concentration, temperature, and photosynthesis rate are interdependent and vary diurnally. The decrease in the DO concentration during nights due to inhibition of photosynthetic activity and raise in DO concentration due to active photosynthesis of microalgae during the day time have been observed (Saba et al., 2017). Studies on algal-biocathode MFCs suggest optimum DO for maximum power output is between 4.5 and 5.5 mg/L (Juang et al., 2012; Rodrigo et al., 2007). Jung et al. used C. vulgaris as an algal-biocathode system to study the influence of temperature and light on DO concentration (Juang et al., 2012). The algal-biocathode system was illuminated using an artificial lamp with varying power from 6 to 26 W. It led to the change in the temperature from 27 C to 31.2 C and increased DO concentration from 3.7 to 6.8 mg/L. Whereas freshwater provided with similar condition serving as control resulted in decreased DO concentration upon an increase in the temperature. A similar type of research carried by other researchers concluded that the performance of algal-biocathode MFC relies on the cathodic biocatalysts and DO concentration (Lobato et al., 2013; Gouveia et al., 2014). Gouveia et al. (2014) and Wu et al. (2013, 2014) also reported a tenfold increment in the microalgal growth rate and sixfold increased power production with increase in DO and chlorophyll concentrations at a high light intensity.

10.4

Conclusion

Both approaches, using photosynthetic microalgae as the algal biocathodes in MFC cathode chambers and coupling MFC with algal photobioreactor system, have been reported to result in enhanced electricity production and efficient pollutant removal from wastewaters. The approach of integrating MFC and algae cultivation systems has been proven to overcome inefficient nutrient removal and mechanical aeration
associated limitations of the MFC technology. An attractive benefit of this integration is the growth of algal biomass that can be harvested for several products. Overall, this approach appears to be economically attractive and sustainable option for wastewater treatment and energy recovery. Further studies need to consider the design aspects along with optimization of biotic and abiotic factors toward upscaling the process. The focus should also be on the selection of suitable microalgal strain for maximizing the oxygenic photosynthesis and lipid production processes in algal-biocathode MFCs. Finally, life cycle and techno-economic assessments need to be performed to elucidate the applicability of the integrated MFC-microalgae cultivation systems.

Acknowledgment P. Chiranjeevi gratefully acknowledges the postdoctoral research fellowship from IISER Mohali.

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2081.

Integration of bioelectrochemical systems with other existing wastewater treatment processes

11

Makarand M. Ghangrekar, Sovik Das and Bikash R. Tiwari Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India

Chapter Outline 11.1 Introduction 229 11.2 Integration of bioelectrochemical system with electro-Fenton process 232 11.3 Integration of bioelectrochemical system with aerobic processes 235 11.4 Integration of microbial fuel cell with anaerobic digestion 240 11.5 Microbial fuel cell integration with septic tank 241 11.6 Microbial fuel cell integration with dark fermentation 242 11.7 Microbial fuel cell integration with microalgae 242 11.8 Novel integration of other processes with microbial fuel cells 243 11.9 Way forward 244 References 245

11.1

Introduction

As human societies have become increasingly industrialized and urbanized, not only the use of natural resources has been skyrocketing but also there has been an enormous increase in the amount of waste materials generated. When the population of the Earth was much smaller (e.g., fewer than 2 billion), the traditional “take, make, waste” pattern of resource consumption was acceptable. However, now recycling and reuse of all types of resources (including water) have become essential, and there must be an increase in the use of renewable resources. In contrast to many other natural resources, water is inherently renewable. Mother Nature has been recycling water since the origin of life on the planet. Recycling technologies can significantly reduce net water abstraction from the environment; however, many of those technologies require an increase in the consumption of other resources, especially energy. In our resource-constrained world, increasing the consumption of any resource, even for necessary functions such as water management, must be carefully considered. Environmental goals include meeting water needs from locally available water supplies while maintaining energy neutrality, minimal chemical consumption, and responsible nutrient management. Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00011-4 © 2020 Elsevier Inc. All rights reserved.

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The organic fraction of waste present in various domestic and industrial wastewaters can be a considerable feedstock, which is yet to be exploited to the fullest extent for recovering renewable and providing a suitable alternative to fossil fuel
based economy. Nevertheless, various treatment options have been utilized to address this issue. Anaerobic digestion can be used to reduce the chemical oxygen demand (COD) by more than 70%, with a biogas recovery of 80%
90% at an industrial scale (Sankaran et al., 2014). The metabolic pathways prevalent in anaerobic digestion are hydrolysis, acidogenesis, acetogenesis, followed by methanogenesis. Upflow anaerobic sludge blanket (UASB) reactor has a critical advantage over other anaerobic digestion configurations in terms of ease of operation at low hydraulic retention time (HRT) and applicable higher organic loading rates (OLRs). The reactor configuration is in upflow mode with the bacterial growth following a suspended growth pattern. The sludge bed volume, which occupies the lower portion of the reactor, is about one-third to half of the reactor volume (Lettinga, 1995). The gas
liquid
solid separator placed at the upper portion of the reactor allows the retention of solids while simultaneously allowing gas to be collected for further flaring or onsite utilization. The mixing of the wastewater is ensured by the upflowing wastewater and the rising gas bubbles (Bhunia and Ghangrekar, 2008a), while the sludge bed at the bottom provides an intricately connected microchannel network for the liquid movement through the sludge, thus facilitating proper contact between the substrate in the effluent and the microorganisms in the granules. Granulation is a major step in the proper functioning of UASB reactor and can be enhanced by the addition of external agents such as cationic polymers (Bhunia and Ghangrekar, 2008b) or by ensuring proper hydrodynamic conditions in the sludge bed of UASB reactor to favor biomass granulation (Bhunia and Ghangrekar, 2008a). While the wastewater treatment technologies based on anaerobic digestion have been popular for the last three decades, bioelectrochemical systems (BESs) are slowly gaining footage with the intent of providing a green and sustainable solution to deal with xenobiotics, pharmaceuticals and metals, and recovering valuable byproducts. Microbial fuel cell (MFC) is an upcoming promising technology, which can convert the chemical energy present in the bonds of organic matter to electrical energy with the help of electroactive microorganisms (Du et al., 2007). The anodic chamber of MFC is maintained under anaerobic conditions, where electrogenic microbes act as the biocatalyst and have the capability to donate electrons to a solid electrode. The MFC can utilize organic matter present in various wastewaters originating from domestic and industrial sources. The anodic half-cell reaction for glucose oxidation is shown in the following equation: Anode: C6 H12 O6 16H2 O ! 6CO2 124H1 124e2 ΔE˚ 520:41 V

(11.1)

Similarly, the cathodic half-cell reaction for oxygen as an electron acceptor can be represented by the following equation: Cathode: O2 1 4H1 1 4e2 ! 2H2 O

ΔE˚ 5 0:82 V

(11.2)

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In the cathodic chamber the electrons and protons combined with the terminal electron acceptor such as oxygen, nitrate, or other chemical oxidants to produce water as mentioned in Eq. (11.2) (Rismani-Yazdi et al., 2008) or other reduced products as per the chemical oxidant used. The cathodic chamber in the MFC can be eliminated in case the cathode is directly exposed to air in an arrangement, also known as single-chamber MFCs. Anaerobic digestion can easily deal with solid waste with high moisture content and thus eliminates the need for moisture reduction prior to treatment. However, MFC technology has the upper hand while dealing with wastewater having a low concentration of organic matter and even at low-temperature conditions, that is, below 20 C where anaerobic digestion faces operational difficulty. The current trends in MFC research hold promise for an MFC-centered treatment scheme as a technically viable and economically feasible option in real-field applications to achieve sustainable wastewater treatment. Still, the energy output achieved in milliliter-scale MFCs is not yet replicated on a large scale owing to the occurrence of overpotential losses during electron/charge transfer, and substrate and electron scavenging by nonelectrogenic microorganisms. Moreover, the higher capital investment required for the fabrication of MFCs can be attributed to the various structural components such as electrodes, membranes, and catalysts (Tiwari et al., 2016, 2017). A significant cost (B$1600 m22) is ascribed to the commonly used Nafion 117 membrane, which is used to separate anodic and cathodic chamber. While the use of low-cost polymeric and ceramic membranes can reduce the cost associated with membranes, a more attractive alternative could be the implementation of membrane-less MFCs; however, the membrane-less MFCs are facing problems, such as high internal resistance, difficulty in maintaining voltage when put in series, and substrate loss, those need to be resolved by properly configuring the system. The effective combination of MFC with anaerobic digestion is expected to make amends for the energy incurred in wastewater treatment and simultaneously reduce the environmental footprint. The integration of anaerobic digestion and MFC can eliminate the need for energy-intensive aerobic treatment, which is generally provided as a posttreatment step to anaerobic treatment. Integration of anaerobic digestion and MFC is energy positive; however, the economic return needs a significant impetus to approve practical implementation (Beegle and Borole, 2018). While dealing with the treatment of high-strength wastewater, MFC can be used in combination with another complementary technology, which can not only aid in reducing the organic matter content but also provide waste to energy conversion. Anaerobic treatment technologies can be selected, since they eliminate the use of aeration, produce less sludge, and can handle higher OLRs. The UASB reactors, which are recognized as one of the most feasible technologies for the treatment of domestic wastewater treatment in developing countries such as Brazil, India, and Colombia, can also be used for first-stage pretreatment of such high-strength wastewater. The valuable by-product methane generated can be recovered and converted into electrical energy (Mendoza et al., 2009). Industrial-scale application, high efficiency, and relatively low cost compared with aerobic technologies are the noticeable advantages of the UASB reactor. Since the UASB and MFC are not competitive but could be complementary to each other,

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they can be used in combination with a more efficient wastewater treatment through bioconversion (Pham et al., 2006). As the cathodic and the anodic chambers of MFC are similar to aerobic and anaerobic reactors, the proposal of integrating MFCs with a UASB reactor could lead to extensive benefits in both organic matter removal and electricity generation (Cheng et al., 2010). BESs are novel and embryonic technologies that can not only treat wastewater but can simultaneously harvest electrical energy and other valuables present in wastewater. They can also be used for the production of hydrogen, nutrient recovery, and desalination of brackish water (Li et al., 2014a). Typically, in BES, electrochemically active microbes degrade organic matter present in wastewater to produce electrons, which are then accepted by the anode and transferred to the cathode through an external circuit. On the cathode the electrons combine with the protons, transported from the anodic chamber through a proton exchange membrane (PEM), and terminal electron acceptors such as oxygen to form water, and thus, bioelectricity is generated in the process. Although BES can significantly reduce the organic content of wastewater, the capital and operation and maintenance costs associated with BES need to be reduced drastically for the successful field-scale demonstration of such a novel technology. This can be done by coupling existing wastewater treatment technologies, which offer wastewater treatment at a relatively lower cost, with expensive BES, thus reducing the overall cost of the system and offering a higher degree of treatment to the wastewater (Yuan and He, 2015). Traditionally used wastewater treatment technologies are not environmentally sustainable as these technologies fail to recover value-added products while simultaneously treating wastewater, which on the other hand, is the major advantage of using BES for the abatement of water pollution. Thus the integration of BES with existing wastewater treatment technologies can be envisioned, where wastewater can be considered as a resource for the production of valuables using BES with the simultaneous treatment of wastewater at a reasonably lower cost (Li et al., 2014b). Aerobic processes such as activated sludge process (ASP) can also be coupled with BES to reduce the organic load on aerobic treatment, thus reducing the cost associated with aeration and simultaneously recover bioelectricity in the process. As anodic chamber of BES function following the anaerobic pathway for the degradation of organic matter present in wastewater, the sludge generation in the process is drastically reduced, thus circumnavigating the problem of large volume of excess sludge management, which is a major concern for traditionally used aerobic processes (Zhang et al., 2013). Only the treatment of wastewater is achieved using ASP, but coupling BESs with ASP can be a feasible option of extracting the energy stored in wastewater in the form of clean bioelectricity.

11.2

Integration of bioelectrochemical system with electro-Fenton process

Electro-Fenton (e-Fenton) process can be employed for the destruction of organic and biorefractory pollutants present in wastewater by highly reactive hydroxyl

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Figure 11.1 MFC coupled with bio-electro-Fenton. MFC, Microbial fuel cell.

radicals produced from the reaction of electrochemically generated H2O2 with Fe21 (Brillas et al., 2009). The electrons required for the production of H2O2 can be supplied from the biotic anode of a BES, which are produced as a result of bioelectrochemical oxidation of organic matter present in the wastewater, thus reducing the overall operational cost of the system (Fig. 11.1). Such a combination of BES with the e-Fenton process is generally termed as the bio-electro-Fenton (BE-Fenton) process, which has been widely used for the degradation of refractory compounds, such as dyes, present in wastewater (Feng et al., 2010). This kind of integration is not only successful in reducing the operating cost of BE-Fenton, but it is also able to harness the chemical energy present in wastewater in the form of combustionless and pollution-free bioelectricity (Rittmann, 2008). The cathodic oxygen reduction reaction (ORR) taking place in a BES can be directed to follow the two-electron pathway with the assistance of a suitable cathode catalyst to produce H2O2. This cathodic H2O2 can not only be used for the removal of recalcitrant and/or emerging pollutants but also for the simultaneous disinfection of the anodic effluent of the MFC. The hydroxyl radicals, thus produced in the cathodic chamber, have also been reported to enhance the power output of an MFC (Zhuang et al., 2010a). The cathodic Fenton reaction in a dualchambered MFC fabricated using expanded polytetrafluoroethylene-laminated cloth as a PEM enhanced the power production by four times. However, such a power was not long lasting as the Fenton reagents were consumed in a short duration of time. This obstacle was overcome by using Fe2O3/carbon felt composite cathode, which resulted in not only higher power generation but also enduring and sustainable power from the modified MFC (Zhuang et al., 2010a). The Fe2O3 present in the composite carbon felt cathode catalyzed Fenton reaction and generated H2O2 by supplying electrons to hydroxyl radicals. The composite cathode produced a maximum power density of 341.4 mW/m2; however, the H2O2 produced in the cathodic chamber was instantaneously consumed due to the presence of ferrous ions.

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Integrated Microbial Fuel Cells for Wastewater Treatment

First, the oxygen was absorbed onto the surface of carbon felt followed by the reduction of Fe2O3 to form ferrous ions. Following the two-electron pathway for ORR, the H2O2 was produced and reacted with Fe21 to produce OH and Fe31. The hydroxyl radicals then reacted on the cathode to form water. Thus the study demonstrated the successful cathodic production of H2O2 applying novel composite cathode coupled with enhanced power recovery in MFC due to the BE-Fenton reaction. The highly active hydroxyl radicals formed through the process of BE-Fenton can also be used to degrade persistent organic pollutants and refractory compounds such as dyes. Such an application of BE-Fenton was demonstrated by Zhuang et al. (2010b), where Rhodamine B (RhB) was decolorized in the cathodic chamber of MFC. An Fe2O3/carbon felt composite cathode produced hydroxyl radicals through e-Fenton reaction, which degraded RhB more efficiently during short-circuit conditions rather than during a closed-circuit condition. Total organic content and RhB were removed by 90% and 95%, respectively, in only 12 h during short-circuit condition. Both total organic matter removal and decolorization efficiency were improved when the MFC was operated in short-circuit mode in comparison to a closed-circuit condition with an external load resistance of 1000 Ω. This was mainly attributed to the higher cathodic current density during short-circuit conditions, which favored the production of H2O2. However, COD removal efficiency was not found to be drastically affected by the external resistance of the MFC. Concomitantly, power generation was also observed for the MFC with steady current and maximum power density of 0.61 mA and 307 mW/m2, respectively, with an external load resistance of 1000 Ω (Zhuang et al., 2010b). Another azo dye, namely, amaranth with the concentration of 75 mg/L, was degraded by 82.59% within 1 h using 1 mmol/L of Fe21 (Fu et al., 2010). Simultaneously, 28.3 W/m3 of power was also recovered in the process. Thus the application of the bio-electronFenton (BEF) reaction in the cathodic chamber of MFC can degrade various azo dyes and simultaneously recover power from the process. Emerging micropollutants such as estrogens, comprising 17β-estradiol (E2) and 17α-ethynyl-estradiol (EE2), were also removed in the cathodic chamber of an MFC employed with an Fe2O3/carbon felt composite cathode. When operated in a closed-circuit condition, 81% of E2 and 56% of EE2 were removed at a retention time of 10 h in the cathodic chamber of MFC. Also, the highest concentration of H2O2 in 10 h reached up to 1.2 mg/L (Xu et al., 2013). Simultaneous power recovery of 4.35 W/m3 was also achieved using the MFC. Thus the study proved that micropollutants such as estrogens could be successfully removed in an MFC, which is designed to undergo BE-Fenton reaction in the cathodic chamber and concomitantly recover power from the process. In another application of BEF reaction in an MFC, paracetamol (PAM), an emerging contaminant was also successfully degraded by 70% within a reaction time of 9 h (Zhang et al., 2015). The highest degradation efficiency of PAM was obtained when an external resistance of 20 Ω was used with the cathodic pH of 2. The degradation of PAM resulted in the formation of intermediate metabolites such as p-nitrophenol via p-aminophenol and to less hazardous dicarboxylic/carboxylic acids. The flux of electrons coming from the anode was found to be positively G

Integration of bioelectrochemical systems with other existing wastewater treatment processes

235

influencing the degradation efficiency of PAM. When the MFC was operated in a closed-circuit mode, an average power density and current density of 217.27 6 23.24 mW/m2 and 757.41 6 65.47 mA/m2, respectively, were reported (Zhang et al., 2015). Thus the successful application of BEF-based MFC for the degradation of emerging contaminants such as PAM was demonstrated. Another emerging contaminant, namely, triphenyltin chloride (TPTC), was also successfully degraded by 78.32% 6 2.07% in an MFC with Fe2O3/graphite felt (GF) composite cathode following BEF reaction (Yong et al., 2017). The TPTC degradation was initiated with the cleavage of Sn
C bonds. The TPTC degradation rate of 0.775 6 0.021 μmol/L/h was found to be significantly higher. This was mainly due to the reaction of highly active hydroxyl radicals formed in the cathodic chamber due to the BEF reaction with TPTC. A maximum of 135.96 μmol/L of H2O2 was generated using the composite cathode of Fe2O3 and GF, which further reacted with Fe21 to produce hydroxyl radicals. Simultaneously, a maximum voltage and maximum power density of 174.4 mV and 57.25 mW/m2, respectively, were also obtained from the MFC that degraded TPTC. Thus the study reaffirmed the fact that emerging pollutants can be successfully degraded in an MFC by following the BEF reaction.

11.3

Integration of bioelectrochemical system with aerobic processes

The aerobic cathodic chamber of BES can be combined with the aeration tank employed in ASP, thus reducing the load on the aerobic process and hence reducing the operating cost by minimizing the aeration required for treating the wastewater (Fig. 11.2). Such integration will also aid in removing particulate pollutants from wastewater, which is generally not removed in BES as the latter is generally considered as an attached growth process, where the biofilm is tethered to the anode (Cha et al., 2010). Thus this kind of integration can easily meet the required discharge quality standard, which has become a major area of concern in yesteryears. However, if the major part of the organic content is removed in the BES part of the combined system, then additional carbon sources may be required to sustain the population of biological consortia in the aerobic region. This problem can be eluded by implementing such systems for the treatment of high-strength wastewater, thus solving the problem of the addition of external carbon sources for the aerobic process. Such integration was successfully demonstrated by Cha et al. (2010), where the aeration tank of ASP was used as the biocathodic chamber of MFC. The oxygen present in the air pumped into the aeration tank of ASP was used as the terminal electron acceptor for the ORR occurring at the cathode of MFC. The combination of ASP and MFC has numerous advantages such as (1) it can be easily fitted in existing plants without major modifications; (2) oxygen can be shared for cathodic reduction in MFC and for the aerobic oxidation of organic matter in the aeration

(A)

(B)

Influent

(C)

Influent Influent

Anodic chamber of MFC

Aeration tank/cathodic chamber

Clarifier

Effluent

Anodic chamber of MFC

MBR

Effluent

Anodic chamber MFC

Air

Aerobic tank/ cathodic chamber

Clarifier

Effluent

Recycled sludge Waste sludge

Waste sludge

Waste sludge

Figure 11.2 (A) MFC combined with solid contact (aeration) tank, (B) MFC combined with MBR, and (C) MFC submerged in the aerobic tank. MFC, Microbial fuel cell; MBR, Membrane bioreactor.

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tank; (3) aerobic biofilm attached to the cathode can aid in ORR and organic matter removal; and (4) the sludge generation can be reduced as the anodic chamber of MFC operates anaerobically, which produces lesser sludge in comparison to ASP. In an innovative demonstration by Cha et al. (2010), membrane-electrode assembly (MEA) built on the concept of single-chambered MFC was developed, and various operating parameters were optimized. Its potential for the successful field-scale integration of MFC and ASP was also evaluated. The MEA with an electrode spacing of 3 mm was used. This was done to minimize the resistance encounter during the transfer of protons between the two electrodes. Numerous combinations of electrode materials, namely, GF and carbon cloth (CC), were also tested to determine the effect of various electrode material on the performance and biofilm growth. The concentration of dissolved oxygen (DO) in the catholyte has an imperative role in the power output of an MFC. Also, too higher DO would increase the operating cost without improving much the performance of the system; thus flow rates of the blower, mixing intensity was varied, and its effect on the cathodic reduction reaction was evaluated. Furthermore, the performance of an MFC submerged into the aeration tank of ASP was evaluated by comparing the performance of an abiotic cathode to abiotic cathode. The integrated system was fed with wastewater with a COD of 0.3 g/L. When different electrode material, namely, CC and GF, were tested, the MFCs with GF anodes demonstrated a higher power density of 17.1 W/m3, and corresponding lower internal resistances of 17 Ω each (Cha et al., 2010). It was mainly attributed to the greater degree of biofilm attachment on GF due to its higher specific surface area compared to CC, which was also confirmed by scanning electron microscopic images. To evaluate the effect of DO of the catholyte on the performance, the blowers were stopped, and the variation in cell potential was noted. The cell potential started to drop rapidly and reached 30 mV, though DO was around 8 mg21. However, the cell voltage was restored almost immediately when the blowers were switched on. Predictably, when the flow rate of air was increased from 1 to 10 L/min, the power density of the MFC increased from 14.7 to 22.2 W/m3. This was mainly due to the higher availability of DO in the catholyte, thus boosting the ORR occurring on the cathode. Even though the performance of biocathode MFC was inferior to that of the MFC with abiotic cathode, biocathodes could be viable alternative both in terms of economic and environmental sustainability as microbes are self-sustainable and are cost-effective when compared to costly metal catalysts. Nevertheless, further investigations should be directed to elucidate on the suitability of biocathodes in MFCs. In a bench-scale study by Gajaraj and Hu (2014), the MFC was incorporated into the modified Ludzack
Ettinger (MLE) process followed by the performance evaluation of such an integration. The bioreactor was constructed with glass having a volume of 7.2 L, and it was divided into three zones, namely, anaerobic/anoxic chamber (far left), aerobic chamber (middle), and an internal settling chamber (far right), separated by plastic baffles. The anodic chamber of the MFC was fitted into the anaerobic/anoxic chamber, whereas the cathodic chamber of the MFC was inserted into the aerobic chamber of MLE. The test MFC was operated in closed-circuit mode

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with external resistance of 1000 Ω, and the control MFC was operated in open-circuit mode. The anodic chamber of the integrated system was inoculated with return activated sludge obtained from the Columbia wastewater treatment plant and it was fed with synthetic wastewater having COD concentration of approximately 500 mg/L, total nitrogen of 51.7 mg/L, 30 mg/L of ammonia
nitrogen (NH41
N), and 6 mg/L of total phosphorus (Gajaraj and Hu, 2014). These closed- and open-circuit systems were operated at an HRT of 1 day with a flow rate and solids retention time of 7.2 L/day and 20 days, respectively. Also, the power generation of the integrated system was monitored regularly to determine the effect of such integration on the power performance of MFC coupled with MLE. The combination of MFC with MLE aided in NO32
N removal efficiency by 31% and reduced sludge generation by 11% as compared to the control and operating voltage of around 0.13 V was also generated from the closed-circuit system (Gajaraj and Hu, 2014). However, no major difference was found when the sludge volume index, bacterial specific oxygen uptake rate, and ammonia-oxidizing bacterial population were determined for the test and control. For the coupled system, more than 97% COD removal efficiency was observed with 537 6 56 mg/L of influent COD. Therefore from the results obtained in this investigation, it can be said that the integration of MFC with MLE could improve the degree of treatment received to the wastewater both in terms of nitrogen and organic matter removal. In another interesting integration, ASP was converted into cassette-electrode MFC (CE-MFC) with the application of aerobic sludge present in the aeration tank as the source of inoculum for the anodic chamber (Yoshizawa et al., 2014). Initially, a laboratory-scale model with a working volume of 1 L was operated in ASP mode with synthetic wastewater having COD of 500 mg/L as the feed, and an HRT of 24 h was adopted. After stable organic matter removal was obtained in the ASP mode, aeration was stopped, and CE-MFC was inserted into the aeration tank to begin the operation of the same. It was found that during the two modes of operation, namely, ASP and CE-MFC, the COD removal efficiency did not change considerably. Also, the power density of 150
200 mW/m2, CE of 20%
30%, and COD removal efficiency of 75%
80% were obtained in CE-MFC. It was proved by Yoshizawa et al. (2014), that aeration tanks present in ASP could be converted into energy recovery system by inserting cassette-electrode assembly. Such integration would also reduce the capital cost associated with the fabrication of a new MFC-based treatment plant. Therefore MFC-based energy-saving and energyproducing wastewater treatment plant can be fabricated without the need for major modifications of the existing treatment systems. However, the cost of membranes and electrodes required for the fabrication of MFCs needs to be reduced considerably for the successful full-scale application of MFC in the wastewater treatment plant. The combination of ASP and single-chambered MFC was also reported to degrade Acid Navy Blue R (ANB) dye with concentration ranging from 50 to 400 ppm (Khan et al., 2015). The synthetic wastewater containing glucose as the carbon source and ANB was initially treated anaerobically in the single-chamber MFC, followed by the degradation of ANB by-products using ASP. The MFC was

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made with a plexiglass chamber with an anodic chamber volume of 100 mL and graphite rods with 15.115 cm2 as electrodes. Nafion was used as the PEM separating the anodic chamber of the MFC and ASP. Anaerobic sludge with mixed liquor suspended solids concentration of 3 g/L was collected from the Okhla sewage treatment plant, New Delhi, India, and used as for the inoculation of the anodic chamber of the MFC. The microbial population residing on the anode was also identified to elucidate the role of different electrogenic species in the degradation of ANB. Exoelectrogenic Geobacter was found in abundance during bacterial quantification, thus confirming its robustness while simultaneously degrading ANB and producing electrons in the process. It was also found that pseudo-first-order reaction kinetics was followed with a negative value of Gibbs free energy during the degradation of the dye. Maximum CE and power density of 10.36% and 2.24 W/m2, respectively, were reported when 200 ppm of ANB was spiked into the synthetic wastewater. Concentration higher than 200 ppm of ANB resulted in lowering the cell potential and thus also reducing the power density of the MFC. This could be attributed to the inhibitory effect of ANB on the microbiota present in the anodic chamber of the MFC. Cyclic voltammetry (CV) studies were carried out, and pronounced redox peaks were found for the whole-cell CV of the MFC, thus proving the proper functioning of MFC when dosed with ANB. After the stabilization of the performance of MFC, almost constant COD removal efficiency of 80% was found for all the dosages of ANB. However, after the successive aerobic treatment, COD removal reached more than 90% for the combined system. Bioelectricity generated with the different dosages of ANB was found to increase with the increase in the concentration of ANB up to 200 ppm, and for higher concentration, decreasing electricity was observed. Power density, calculated against 470 Ω of external resistance, was found to be inversely proportional to the COD removal in the anodic chamber with the maxima again being at the dosage of 200 ppm of ANB (Khan et al., 2015). During the degradation of ANB in the MFC, it was transformed into 1-naphthalenamine, Broener’s acid, and aniline. These compounds were further degraded to phthalic acid and its derivative diethyl phthalate in the ASP employed after MFC. Therefore the integration proved potent enough to successfully degrade dyes, which are typically not removed in traditionally used technologies. Also, it concurrently generated power in the form of bioelectricity in the process. Thus such an integration can be envisioned for the removal of refractory compounds such as dyes generally present in industrial wastewater. Organic matter removal, sludge production, and electricity generation were compared to the integration of MFC with ASP treating synthetic domestic and synthetic industrial wastewater by Asai et al. (2017). An ASP with 1.5 L of aeration tank was combined with 1.5 L of CE-MFC equipped with six CEs. The electrodes were made of GF, each with a surface area of 126 cm2. Return sludge obtained from Asakawa Water Reclamation Center in Tokyo, Japan, was used as the source of inoculum for the ASP portion of the integrated system. Nearly similar COD removal efficiency of about 93% and 97% were obtained, treating synthetic domestic and synthetic industrial wastewater, respectively (Asai et al., 2017). The sludge generation for the CE-MFC was quantified, and it was found that the MFC generated around one-third

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or lesser sludge in comparison to the ASP mainly due to the anaerobic mode of growth followed by the microbes present in the anodic chamber of the MFC. The continuous current generation was also observed during the treatment of two different waste streams with CE greater than 20% for the CE-MFC. It also revealed an added advantage of lesser sludge production when using anaerobic technologies such as MFC, thus reducing the cost associated with the treatment of waste sludge. Thus the fact that MFC, coupled with ASP, can be employed not only for the treatment of domestic wastewater but also for high-strength industrial wastewater is emphasized with the advantage of simultaneous bioelectricity recovery.

11.4

Integration of microbial fuel cell with anaerobic digestion

Combinations of treatment technologies were also employed for achieving higher degradation of distillery spent wash. It is a known fact that sulfide is generated during anaerobic digestion, which can add to the cost of the total system owing to the installation of dedicated treatment units for sulfide removal. MFCs were coupled with UASB to address this issue (Rabaey et al., 2006). The UASB reactor achieved an acetate removal of 78%, while the MFC could successfully remove the sulfide and acetate present in UASB effluent by 98% and 46%, respectively. This results in evidence that MFCs can be employed as a polishing treatment for effluents generated from anaerobic digesters. An MFC can also be inserted into the anaerobic tank of existing wastewater treatment, which would enhance the efficiency of the system (Fig. 11.3). The effluent of the anaerobic zone was fed to a submersible MFC (SMFC), where domestic wastewater served as both as the substrate and source of inoculum (Min and Angelidaki, 2008). The SMFC could achieve a maximum power density of 204 mW/m2 (current density of 595 mA/m2). The high internal resistance

Figure 11.3 Cathodic chamber of MFC inserted into anaerobic tank working as the anodic chamber of MFC. MFC, Microbial fuel cell.

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and the inefficient cathodic electron were found to be the major limiting factors. Hence, further fieldscale studies based on these results could be conducted to access the potential for electricity production from existing anaerobic processes. A single-chamber MFC was coupled to a hydrogen-producing biofermenter to optimize the energy generation from the integrated system in terms of electricity generation and hydrogen production (Sharma and Li, 2010). The specific hydrogen yield of the biofermenter was found to increase with the reduction in OLR, and at the lowest OLR of 4 g/L  day, the hydrogen yield of 2.72 mol H2/mol of glucose was achieved. When the effluent of the biofermenter was fed to the MFC, the highest power density of 4200 mW/m3 was obtained at a coulombic efficiency of 5.3%. Moreover, the COD removal of the system was improved to 71%, and the energy conversion efficiency was enhanced to 29%. When molasses wastewater having COD of 127,500 mg/L was used as a substrate, an integrated UASB
MFC biological aerated filter system demonstrated overall COD removal of 53% and a power density of 1410.2 mW/m2 in the MFC (Zhang et al., 2009). In addition, the hybrid system also achieved a sulfate and color removal efficiencies of 53.2% and 52.7%, respectively. The UASB reactor dominated the COD removal and sulfate reduction, while the MFC unit generated electricity from the oxidation of generated sulfide, and the biological aerated filter contributed majorly to color removal and phenol derivatives degradation. In a UASB-MFC
combined system the COD/sulfate ratio of 3.7 and HRT of 55.6 h are reported to be the optimum condition for maximizing total sulfate removal efficiency (Zhang et al., 2012). However, in order to achieve maximum power output, a COD/sulfate ratio of 2.3 and HRT of 54.3 h are suggested (Zhang et al., 2012). An integrated system of UASB and MFC was employed to treat oilfield wastewater, and performance was evaluated under three different HRTs of 6, 13, and 20 h (Gong and Qin, 2012). It was found that optimal HRT of 20 h was suitable for achieving maximum COD and NH3
N removal efficiencies of 93.5% and 83.4%, respectively. However, the MFC demonstrated a maximum power density of 93 mW/m2 at an optimum HRT of 13 h.

11.5

Microbial fuel cell integration with septic tank

A pilot-scale three-chambered MFC-based latrine was established in Ghana to simultaneously carry out organic matter removal in the anodic chamber, nitrification in the middle chamber, and a biocathode in the third chamber facilitates denitrification (Castro, 2014). Maximum COD removal of 90%, nitrate removal of 76.8%, and power density of 3.4 nW/m2 were achieved during operation. The solid waste was eventually used for composting. An easily pluggable three-column MFCstacked configuration was designed, which can be plugged in a septic tank to generate a power density of 142 mW/m2 (Yazdi et al., 2015). This system is expected to a daily power of 24 W h, which can power a 6-W LED bulb for 4 h and would cost around $25. Similarly, a set of 15 MFCs was stacked in a septic tank of working

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Integrated Microbial Fuel Cells for Wastewater Treatment

volume 2.44 m3 and connected to a power-management circuit to eventually discharge a maximum current of 1.98 mA and power of 4.51 mW (Alzate-Gaviria et al., 2016).

11.6

Microbial fuel cell integration with dark fermentation

The waste by-product generated during biodiesel production, that is, crude glycerol, could be converted to bioenergy by a hybrid system of dark fermentation, MFC, and microbial electrolysis cell (MEC). This integrated approach serves to overcome the thermodynamic limitation faced in dark fermentation, wherein the complete degradation of crude glycerol is difficult (Keskin et al., 2011). The effluent of the dark fermenter still contains high proportions of unutilized organic metabolites ¨ zkan et al., 2012), which can be either converted to electricity in MFCs or hydro(O gen in MECs. It was found that when crude glycerol having an initial COD concentration of 7610 mg/L was utilized in dark fermentation, maximum H2 production of 332 mL/L (yield of 0.55 mol H2/mol glycerol) was achieved (Chookaew et al., 2014). Subsequently, the effluent after 50% dilution was degraded in an MFC to achieve a power density of 92 mW/m2 and a COD removal of 49%. In an alternate approach the diluted effluent was fed to MEC supplied with an external voltage of 1.0 V, which demonstrated an H2 yield of 106 mL/g COD consumed [H2 production rate of 0.05 m3/(m3 day)]. A combination of MEC powered by MFC and dark fermentation achieved a hydrogen gas production of 14.3 mmol/g cellulose at a rate of 0.24 m3/m3  day while utilizing cellulose as a substrate (Wang et al., 2011) using 50 mL of MEC. The MEC was supplied with a potential of around 0.43 V from two MFCs connected in series. The MEC consumed the effluent of the fermentation unit as the feed and contributed a hydrogen production of 0.48 m3/m3  day (yield of 33.2 mmol H2/g of COD removed) with 72 mL of MEC. It is also of prime importance to note that no external electricity input was utilized for the operation of MEC, which eventually resulted in an energy recovery efficiency of 23% (based on cellulose removed). The scope of this technology can also be extended by the use of other biomass sources such as wastewater from domestic and industrial sources. In addition, the hydrogen production from the MEC can be further improved by increasing the applied voltage by the addition of more MFC units and arranging in stacks.

11.7

Microbial fuel cell integration with microalgae

The development of microalgae-MFC is based on extensive studies conducted on MFC and microalgae (Chen et al., 2011; Mendoza et al., 2015). The integration of algae and MFCs will serve as a novel technology, which utilizes the metabolic reactions of photosynthetic microorganisms to convert solar energy to electrical

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energy (Bombelli et al., 2011). For example, consider the case of a microalgae photo-bioreactor (PHB) where the algae having unicellular chlorophyll are capable of converting carbon dioxide to biomass and oxygen with the help of incident light (Lee et al., 2015). Moreover, the oxygen released during photosynthesis by the Chlorella vulgaris in the cathodic chamber of MFC could serve as the terminal electron acceptor (Kakarla and Min, 2014; Powell et al., 2009). Hence, the integrated system can serve the purpose of carbon dioxide sequestration, oxygen generation, and nitrogen removal from the wastewater (Wang et al., 2010). In an alternate approach the dead microalgae served as the substrate in the anodic chamber of MFCs, while the cathode chamber was inoculated with live algae (Kondaveeti et al., 2014; Rashid et al., 2013). The power generation of a single-chambered photosynthetic MFC was observed where the conductive cathode was fixed with different strains of algae, that is, C. vulgaris, Dunaliella tertiolecta, and Synechocystis sp. The highest power density of 10.3 mW/m2 was observed in MFC having Synechococcus sp. when illuminated with 10 W/m2 of white light. A series connection of four MFCs can generate enough power (potential of 2 V and current around 10 mA) to operate a commercially available small digital clock. An anoxic MFC was developed where the photo-biocathode was illuminated, and dissolved carbon dioxide was used as an electron acceptor. This biocathode successfully fixed carbon dioxide, and the MFC generated a power density of 750 mW/m2 (Cao et al., 2009). An MFC was coupled with a tubular PHB where the PHB served as the cathode compartment inoculated with C. vulgaris (Wu et al., 2013). The oxygen generated by C. vulgaris and power output in MFC were lightdependent. The MFC with algae biocathode achieved a maximum power density of 24.4 mW/m2 under intermittent illumination, which was 2.8-fold higher than that of MFC with an abiotic cathode.

11.8

Novel integration of other processes with microbial fuel cells

Distillery spent wash at various concentrations ranging from 1100 to 10,100 mg/L was first treated in an MFC (Anupama et al., 2013). The COD removal efficiency, as well as power output, increased with an increase in the concentration of substrate till 6100 mg/L. The MFC demonstrated a maximum power density of 18.35 mW/m2 and COD removal of 64% at a feed concentration of 6100 mg/L. The effluent of MFC was further subjected to treatment in a rotating biological contactor, which achieved a COD, BOD, and dissolved solid removal of 84%, 81%, and 46.4%, respectively. The two-stage integrated process of fungal treatment of cereal-based distillery stillage followed by MFC demonstrated an overall COD removal of 99% along with almost complete removal of suspended solids (Ghosh Ray and Ghangrekar, 2015). Chitosan (0.7 g/L of settled sludge) was also obtained as a value-added product. Prior treatment with fungal strain improved the power generation by around

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Integrated Microbial Fuel Cells for Wastewater Treatment

threefold to 2.6 W/m3 (current density of 246 mA/m2) and achieved a 92% soluble COD reduction in MFC operated at an OLR of 1.5 kg COD/m3  day.

11.9

Way forward

MFC is undoubtedly a promising technology for addressing sustainable goals, which can significantly change the wastewater treatment scenario. However, at present when MFCs are limited by low power densities, high fabrication cost, and high operation cost, it seems a better strategy to integrate MFCs with the conventional wastewater treatment technologies to overcome the shortcomings of the later, such as removal of recalcitrant compounds, endocrine disruptors, aromatic compounds, and pharmaceuticals. Unlike conventional wastewater treatment technologies, which have been thoroughly evaluated to achieve superior performance and operational stability by considerable investigation and optimization of various process parameters, MFC technology is still not matured. In addition, in an integrated system, it is necessary to ensure the proper functioning of both the participating technologies, hence adequate attention should be bestowed on understanding the complexities of the integrated system and process control. While dealing with multiple stacks of MFCs during scaling-up, challenges such as short-circuiting and voltage reversal need to be addressed. It is agreeable that supercapacitor-based power storage devices should be encouraged to be used in conjugation with MFCs to improve the utility of the electricity harvested. There has been significant progress and technological improvement in the field of MFC for wastewater treatment. The potential benefits that MFCs offer in comparison to the conventional wastewater treatment system are their sustainable and carbon-neutral approach toward biodegradation of waste and the simultaneous generation of a clean and green source of energy. The electricity generation is a singlestep process, and the overall operation of the process is stable and robust due to the self-regeneration of the anodic microbiota, which offers good resistance to environmental stress. However, performance and cost-related issues with regards to the cathode catalyst, PEM, and upscaling need to be addressed in order to make this upcoming technology a practical solution in the real field. Nonnoble metal catalysts need to be developed for both anode and cathode, which can reduce the corresponding activation overpotential and subsequently improve the electrode performance. Advancement in the MFC architecture is required to ensure maximum contact between the substrate and the biofilm and higher electrode surface to electrolyte volume ratio. The cost of fabrication needs to be reduced in order to commercialize the MFC and compete with the existing wastewater treatment systems. This is only possible when low-cost alternatives for cathode catalyst and PEM are incorporated, which are extremely efficient and support higher current densities. This will allow MFCs to extract maximum energy from the organic matter present in wastewater in the form of electricity. Nonetheless, if MFCs are associated with decentralized wastewater treatment plants, the energy generated from MFCs will reduce the

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operating energy need of the treatment plant. This chapter is expected to discover innovative avenues for the application of MFC centered hybrid systems and expand the horizon of MFCs beyond their present applications.

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23094. Wang, X., Feng, Y., Liu, J., Lee, H., Li, C., Li, N., et al., 2010. Sequestration of CO2 discharged from anode by algal cathode in microbial carbon capture cells (MCCs). Biosens. Bioelectron. 25 (12), 2639
2643. Wang, A., Sun, D., Cao, G., Wang, H., Ren, N., Wu, W.M., et al., 2011. Integrated hydrogen production process from cellulose by combining dark fermentation, microbial fuel cells, and a microbial electrolysis cell. Bioresour. Technol. 102 (5), 4137
4143. Wu, X.Y., Song, T.S., Zhu, X.J., Wei, P., Zhou, C.C., 2013. Construction and operation of microbial fuel cell with Chlorella vulgaris biocathode for electricity generation. Appl. Biochem. Biotechnol. 171 (8), 2082
2092. Xu, N., Zhang, Y., Tao, H., Zhou, S., Zeng, Y., 2013. Bio-electro-Fenton system for enhanced estrogens degradation. Bioresour. Technol. 138, 136
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263. Yong, X.Y., Gu, D.Y., Wu, Y.D., Yan, Z.Y., Zhou, J., Wu, X.Y., et al., 2017. Bio-electronFenton (BEF) process driven by microbial fuel cells for triphenyltin chloride (TPTC) degradation. J. Hazard. Mater. 324, 178
183. Yoshizawa, T., Miyahara, M., Kouzuma, A., Watanabe, K., 2014. Conversion of activatedsludge reactors to microbial fuel cells for wastewater treatment coupled to electricity generation. J. Biosci. Bioeng. 118 (5), 533
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Yuan, H., He, Z., 2015. Integrating membrane filtration into bioelectrochemical systems as next generation energy-efficient wastewater treatment technologies for water reclamation: a review. Bioresour. Technol. 195, 202
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7. Zhang, F., Ge, Z., Grimaud, J., Hurst, J., He, Z., 2013. In situ investigation of tubular microbial fuel cells deployed in an aeration tank at a municipal wastewater treatment plant. Bioresour. Technol. 136, 316
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163.

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12

M.M. Ghangrekar1, G.D. Bhowmick2 and S.M. Sathe1 1 Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India, 2Department of Agricultural and Food Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India

Chapter Outline 12.1 Introduction

249

12.1.1 Wastewater and its sources 249 12.1.2 Conventional wastewater treatment practices and lacunas 250

12.2 Bioelectrochemical systems

251

12.2.1 Evolution tree of bioelectrochemical systems 251 12.2.2 Major forms of bioelectrochemical systems 252

12.3 Membrane bioreactor 252 12.4 Hybrid bioelectrochemical system membrane bioreactor systems: principle, treatment efficiency, and performance index 254 12.4.1 Integrated bioelectrochemical system membrane bioreactor systems 255 12.4.2 Combined bioelectrochemical system membrane bioreactor system 259

12.5 Outlook and future perspectives 12.5.1 12.5.2 12.5.3 12.5.4

262

Water-energy nexus 262 Membrane fouling mitigation 262 Control of emerging contaminants 262 Field-scale applications 267

12.6 Conclusion 267 Acknowledgment 267 References 268

12.1

Introduction

12.1.1 Wastewater and its sources A drastic increase in the world population has led to intense industrialization and subsequent wastewater generation. The residential areas also contribute to the pollution by discharging untreated or partially treated sewage in natural water bodies. Sewage is the waste produced by toilets, bathing, laundry, or culinary operations or the floor drains associated with these sources and includes household cleaners and medications apart from the organic matter, suspended solids, and pathogens. Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00012-6 © 2020 Elsevier Inc. All rights reserved.

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Untreated sewage contains readily biodegradable organics, inorganic and organic chemicals, nutrients, suspended solids, and emerging contaminants in the form of heavy metals, detergents, pesticides, fertilizers, antibiotics, etc. which if released untreated in a water body can damage the whole dependent ecosystem. The industrial wastewater generated from different industries contains a heavy load of chemical oxygen demand (COD), biochemical oxygen demand (BOD), heavy metals, etc., which also demands an affordable and reliable treatment system for safe discharge of treated effluent in water bodies or reusing it back to process. Such wastewater needs to be properly treated up to the discharge standards as recommended by water regulatory bodies across the globe before releasing into the water bodies.

12.1.2 Conventional wastewater treatment practices and lacunas Wastewater treatment is closely related to the standards and/or expectations set for the effluent quality by the controlling authority, and treatment plant is designed to meet this set of norms. Primary treatment is provided for the removal of floating solids (screen chamber), oil and grease (skimming tank), suspended inorganic matter if present in wastewater (grit chamber), and primary sedimentation tank for settling of settleable organic solids. Biological treatments are proved to be most reliable for treatment of organic matter present in wastewater, which escape primary treatment. In biological wastewater treatment, dissolved and particulate organic matters present in the wastewater are oxidized. This is achieved by enriching microbes for consuming the organic matter as food and converting it to carbon dioxide, water, and energy for their own growth and reproduction. The biological reactor is then followed by additional settling tank to settle the bacterial cells produced in biological reactor. Anaerobic processes such as upflow anaerobic sludge blanket (UASB) reactor, anaerobic filters are commonly used anaerobic wastewater treatment systems that can provide sufficient treatment efficiency along with bioenergy recovery and less excess sludge generation; however, anaerobic process requires skilled supervision because the anaerobes are more sensitive for changes in environmental parameters and are slow growing. On the other hand, aerobic bacteria have wider tolerance for the changes in environmental conditions; however, aerobic treatment requires higher operating cost because of necessity of aeration or larger area to expose more surface of wastewater to atmospheric air. If wastewater is effectively treated, it can reduce the stress on freshwater by making water available for various uses. For removal of carbonaceous and nitrogenous matter present in the municipal wastewater by employing aerobic processes, the energy requirement of 1.66 and 0.30 kW h/m3, respectively, is reported (Scherson and Criddle, 2014). The aerobic process such as activated sludge process requires B0.6 kW h of energy to treat per m3 of wastewater (McCarty et al., 2011). On the other hand, wastewater generally contains energy potential value of as high as 17.8 28.7 kJ/g of COD (Heidrich et al., 2011). Anaerobic wastewater treatment technologies are capable of harnessing this energy in the form of biogas; however, some of the biogas generated is

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getting lost in dissolved form in the effluent because of higher partial pressure of biogas inside the anaerobic reactor (Smith et al., 2013). Until recent times, the main focus of wastewater treatment is on the removal of suspended solids, biodegradable organic matter, and nutrients; however, with an increase in the use of inorganic and organic chemicals, now there is equal importance on controlling the emerging contaminants. As conventional treatments are not designed for removal of emerging contaminants, it leads to their discharge in receiving water bodies, which can be further detrimental (Petrie et al., 2014). Therefore the development of energy efficient, environmentally sustainable wastewater treatment technologies is the need of the hour to address both water- and energy-related crises.

12.2

Bioelectrochemical systems

12.2.1 Evolution tree of bioelectrochemical systems Organic ingredients present in the wastewater, considered as “misplaced resource,” are being generally treated by biological treatment systems worldwide. The byproducts of these are predominantly the biogas such as methane (CH4) that has further face value in terms of heat or converting this energy in electrical energy. Whereas hydrogen is another derivative of anaerobic digestion systems, which eventually has much higher gross calorific value than the other components of biogas. This fact fostered researchers’ concern on developing processes that proliferate the proton (H1) production, and subsequent bioelectricity generation in a special type of wastewater treatment module-cum-bioelectrochemical system (BES) called microbial fuel cell (MFC) using microorganism as biocatalyst to use wastewater as a fuel source. The BES as we see it today has been incepted a century back by scientist M.C. Potter. Potter demonstrated that the generation of electromotive force from the living culture of Escherichia coli and Saccharomyces cerevisiae with platinum electrodes was attributed to the “fermentative action of yeast” (Potter, 1911). In the 1990s Allen and Bennetto (1993) worked on the application of synthetic mediators in MFC, which leads to the development of so-called analytical MFC that is still in use in almost in the same form to date. On the verge of the last century, Kim et al. (1999) started experimenting on some of the electrochemically active species of bacteria that do not need an external mediator to transport electrons toward the surface of the electrodes in mediator-less MFCs. It drastically reduced the cost of operation of the MFCs by eliminating the need for expensive external mediators. Subsequent to this, the last two decades has witnessed sizeable amount of work done on MFC design, their physical, chemical, and biological operating conditions, optimization of microbial metabolism, proper selection of microorganism, construction materials, etc. to enhance the electron transport vis-a´-vis enhanced the wastewater treatment efficiency and electricity harvesting efficacy of the MFC (Logan et al., 2006; Santoro et al., 2017; Gude, 2016). The MFC technology has gone so far from its infant stage to an altitude, where the power production, as well as

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wastewater treatment efficacy, has been improved by a number of orders of magnitude. However, still many bottlenecks are there that need to be addressed in order to achieve its full-scale field applications in the near future, which will be further elucidated in the consequent sections.

12.2.2 Major forms of bioelectrochemical systems In any BES the anodic oxidation and cathodic reduction reaction cause a redox potential difference that is responsible to force the electrons to flow against the potential gradient, spontaneously. When different kinds of microbes involved in any of these reactions, the system is called microbial electrochemical systems (MXS). One such MXS like MFC uses microbes on its electrodes to create a low redox potential at the anodic side and high redox potential related to oxygen reduction reaction (ORR) at the cathodic side, resulting in a bioelectricity generation. In the case of microbial electrolysis cells (MECs), valuable chemical products are produced by means of applied external potential subject to its cathodic counterpart. When the product is other than hydrogen, namely, any significant complex organic molecules by cathodic reduction of CO2 and any other simple organic molecules present in waste products, those MECs are termed microbial electrosynthesis cell (MES). Desalination can also be achieved using slightly modified MFC design by involving different ion-exchange membranes, which then termed as microbial desalination cell (MDC). Sediment MFC (SMFC) can be used for sediment remediation as well as to generate electricity from redox gradients occurring across the sediment water interface. The design of SMFC is privileged by the in situ applications of MFC without the need of any proton exchange membrane (PEM). This variant of BES is already ready-to-scalable technology with minor modifications, unlike other BES variants, which need substantial research prior to scaling up. Sediment microbial carbon-capture cell (SMCC) is a modified SMFC that uses the algae developed in the water overlying the sediment and considered to be a system capable enough to generate power along with the algal cultivation for biofuel production. The microalgae production integrated with MFC appears to be a promising synergistic approach since it can act as in situ O2 producers, which further endorses the bioelectrochemical reactions in MFC and termed photosynthetic MFC (PMFC). When the enzyme is used to oxidize the “fuel,” that is, the organic matter present in the wastewater or to reduce the terminal electron acceptor in MFCs as biocatalyst, it is named enzymatic fuel cell (EFC). Moreover, BES can also be used as an efficient system for treating the recalcitrant pollutants and toxic wastewaters by means of microbial electro-remediation cells (MRCs) (Fig. 12.1).

12.3

Membrane bioreactor

The membrane bioreactor (MBR) is a suspended growth process with the combination of a membrane filtration, such as microfiltration (MF) or ultrafiltration (UF), and it is now widely used for municipal and industrial wastewater treatment with

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Enzymatic fuel cell (EFC)

Microbial carboncapture cell (MCC)

Sediment microbial Photosynthetic carbon-capture cell microbial fuel cell (SMCC) (PMFC)

Bioelectrochemical systems

Microbial fuel cell (MFC) Sediment microbial fuel cell (SMFC)

Microbial electrolysis cell (MEC)

Hydrogen production

Microbial electrosynthesis (MES)

Microbial desalination cell (MDC)

Figure 12.1 Major forms of bioelectrochemical systems.

plant sizes up to 80,000 population equivalent (i.e., 48 million liters per day) (Judd, 2008). While treating domestic wastewater, MBR evidences higher treatment efficiency; it produces better quality effluent that can be discharged to the coastal, surface, or brackish waterways or to be reclaimed for urban irrigation. Other advantages of MBR over conventional processes include small footprint and easy retrofit to any old wastewater treatment plants. In the past decades, MBR has gained worldwide attention and popularity due to its high treatment efficiency, low excess sludge production, and reliable effluent quality (Williams and Pirbazari, 2007). The aerobic MBR actually combines biological removal of nitrogen, phosphorus, and carbonaceous biological oxygen demand (cBOD), by combining the activated sludge process with membrane filtration, either in submerged or sidestream position, thus producing a particulate-free and reusablequality effluent. Enhanced biological phosphorus (P) removal (EBPR) is also realized, as phosphorus-accumulating organisms release phosphorus under anaerobic conditions and then perform “luxury uptake” of phosphorus under vigorously aerated conditions. Effective physical screening of bacteria and solids by the membranes allows for retention of biologically captured phosphorus as well as total suspended solid (TSS) and particulate cBOD5 in the MBR. Further, adsorption of heavy metals by microorganisms in the activated sludge reduces metal concentration in the treated water (Judd and Judd, 2006). The filtration module used in MBR ensures higher solids retention efficiency as compared to secondary clarifiers and also eliminates the risk of upset of performance due to sludge bulking and filamentous bacterial growth. Like other membrane processes, the operation of MBR is delimited by the transmembrane pressure (TMP). Solid retention on the membrane surface causes fouling and adds up to the hydraulic head, thus causing an increase in TMP values. In order to overcome additional head generated by membrane fouling, more energy cost is incurred. Placement of the membrane module plays an important role in the reduction of the degree of biofouling of the membrane surface (Le-Clech et al., 2005). Membrane placement can be within the bioreactor or as an external unit.

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The external unit as a sidestream operation increases the cost of operation as high energy cost is involved to counter the TMP. In case of the micro-filter module embedded within the system, coarse bubble aeration of aerobic MBR can mitigate fouling by agitation and scouring leading to dislodgment of the deposited particles. Aeration also ensures proper mixing, enhanced biodegradation, and suspension of the coarser particles. Dislodgement of deposited particles and aid in suspension are MBR-specific advantages of the submerged system. In comparison to the sidestream MBR, the submerged MBR system reduces the biofouling mitigation cost by directly utilizing the inbuilt aeration system or tailored to the needs of the membrane surface scouring. The MBRs have been successfully operated in the niche and industrial applications with inlet water quality characteristics a typical of those processed in publically owned treatment works (POTW), including high strength industrial wastewaters from automotive manufacturing, winery and tannery operations, food processing facilities, pig and dairy farming operations, and pharmaceutical manufacturers. However, there are still some hurdles to overcome before its more widespread application, such as high costs of membrane materials, severe membrane fouling, and high energy consumption for aeration (Sharrer et al., 2010). Utilization of cheap, coarse-pore mesh, as an alternative to conventional MF/UF membrane, may lower the construction cost of MBR, increase the economic viability, and promote the application of such processes. Expectedly, this process may go a step forward if the energy contained in wastewater can be recovered to partially offset the energy consumption for aeration (Chu et al., 2008). Anaerobic MBR (AnMBR) proves to be advantageous over conventional anaerobic processes as the land footprint required is less, and complete retention of biosolids can be achieved in the AnMBR system (Lin et al., 2013). The membrane module in case of AnMBR is either MF or UF membranes, while the membrane materials used commonly are ceramic, metallic, and/or polymeric. Ceramic membranes have higher flux rates, as high as 200 250 L/m2/h, as compared to the other two options, and they offer high resistance to biofouling and corrosion (Ersu and Ong, 2008; Ghyoot and Verstraete, 1997). Metallic membranes offer high strength and durability along with high resistance to oxidation and better hydraulic performances. Metallic membranes are also easier to clean as compared to the ceramic and polymeric membranes. However, the cost of ceramic membranes, as well as metallic membranes, is higher than polymeric membranes, which has established polymeric membranes as popular choices in the last two decades (Lin et al., 2013).

12.4

Hybrid bioelectrochemical system membrane bioreactor systems: principle, treatment efficiency, and performance index

An integration of BES with conventional activated sludge process was first reported by Cha et al. (2010) and Min and Angelidaki (2008). In these systems,

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255

Figure 12.2 Venn diagram showing how hybrid BES MBR system interconnected with biological, membrane filtration, and BESs and their pros and cons. BES, Bioelectrochemical system; MBR, membrane bioreactor.

the aeration tank of activated sludge process was directly used as the cathodic chamber, where the aerobic biofilm developed on the cathode served as a lowcost and self-sustainable catalyst. The treatment in aeration tank was followed by a clarifier to support a continuous-flow operation, and excess sludge was returned in succession. Nevertheless, this setup produced extra cost for the clarifier construction. It had been hypothesized that a combined BES MBR system might offer an attractive option for treatment of wastewater. The incorporation of MBR may overcome the drawbacks of BES by improving biomass retention and COD removal efficiency, while BES may generate power to partially offset the energy demand for aeration and filtration in MBR and lower the overall oxygen consumption for COD removal. However, effective integration of these two reactors is still a challenge for pilot-scale applications. It has been proposed that a BES can be connected with an MBR as a pre- or posttreatment step or be directly immersed into bioreactors to recover electricity from wastewater (Yuan and He, 2015; Logan, 2009) (Fig. 12.2).

12.4.1 Integrated bioelectrochemical system membrane bioreactor systems 12.4.1.1 The membrane as cathode-cum-filtration unit A novel bioelectrochemical MBR (BE-MBR), which takes advantage of both MBR and MFC process, was successfully reported by Wang et al. (2011) for the first time. The BE-MBR achieved a maximum power density of 4.35 W/m3 and current

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density of 18.32 A/m3 over 100 Ω of external resistance at a hydraulic retention time (HRT) of 150 min. The coulombic efficiency (CE) was 8.2% along with 92.4% COD and 95.6% NH4 N removal efficiency, and the treated water was almost devoid of suspended solids that attributed to the high biomass retention and solid rejection of the system. Nevertheless, this system has a unique and complex reactor design. Specifically, a stainless steel mesh was used, which played a dual function of filter and cathode. A follow-up investigation by the same group of scientists was done on in situ tubular BE-MBR system for utilization of the generated electricity for fouling control (Wang et al., 2012). Wastewater was pumped into the anodic chamber that was filled with granular graphite and a graphite rod. Effluent from the anodic chamber was fed into the cathodic chamber after passing through the stainless steel mesh, and it was eventually discharged. Performance results of this system showed that the electrical fields formed, which lead to a reduction in the deposition of sludge on the membrane surface. The H2O2 produced at the cathode also contributed to the fouling mitigation by removing the membrane foulants. The maximum power density of 1.43 W/m3 and a current density of 18.49 A/m3 were obtained along with COD removal efficiency of 93.7% and ammonical nitrogen removal efficiency of 96.5%. The developed system has provided a way to suppress the membrane fouling at in situ conditions (Wang et al., 2012). The same mechanism of mitigating membrane biofouling was concluded by Xu et al. (2015) using anthraquinone disulfonate/polypyrrole composite modified polyester flat membrane serving as the cathode of a dual-chamber coupled MFC MBR system. In another investigation, Malaeb et al. used an electrically conductive UF membrane as both the air biocathode and the membrane unit for filtration. The integrated MFC MBR system developed in this experiment used an air biocathode that showed to have good power performance relative to an otherwise identical cathode containing a platinum catalyst. The graphite brush anode was positioned at the center of the integrated MFC MBR system. The permeate quality was comparable to that of a conventional MBR, with removal efficiency for COD, NH3 N and total bacteria of 97%, 97%, and 91%, respectively (Malaeb et al., 2013). The negative charge accumulated on the membrane surface enhanced electrostatic repulsion force between the membrane and foulants, which could reduce biofouling in an experiment by Liu et al. (2014) using membrane as a cathode in MFC-integrated MBR system, although the power output was compromised in this case. The membrane surface to the membrane foulant electrostatic repulsion force was calculated in the range of 2.5 3 10214 N in that integrated system. An anaerobic membrane bioelectrochemical reactor (AnMBER) consisting of hollow-fiber MF membranes with a pore size of 0.4 μm inserted between stainless steel framed cathode was developed by Tian et al. (2014). The system continuously generated bioelectricity of 0.132 V, with a maximum power density of 1.16 W/m3 with 91.6% and 94.8% removal efficiency for COD and nitrate, respectively. Moreover, compared to AnMBR (with open circuit), membrane fouling was mitigated significantly in the AnMBER system that was mainly attributed the reduction in particle zeta potential, favoring agglomeration, and lower amount of

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soluble microbial products (SMP) in the cathodic mixed liquor, which states that BES pretreatment can serve as a way to minimize membrane biofouling in MBR (Tian et al., 2014). The cathode-membrane composite material with a catalyst coated on it was developed by researchers to improve further the power performance of the integrated system. Li et al. developed a cathode membrane with a carbon fiber cloth, polyvinylidene fluoride (PVDF) and C, Mn, Fe, and O consisting catalyst using impregnation and high-temperature pyrolysis. A horizontal distance of 2 cm between the anode and cathode membrane was reported to be sufficient to effectively isolate the environment of the anodic chamber (DO of 0.01 mg/L) from the aerated cathodic chamber (DO of 4.81 mg/L). The performance in terms of COD, NH4 N, and total phosphorus removal efficiency of 90%, 80%, and 65%, respectively, was reported with a catalyst-coated system along with a maximum power density of 1358 mW/m3 (Li et al., 2015b). Filtrable cathode membrane of polyester filter modified with polyaniline (PANI) phytic acid (PA) with Fe Co catalyst on carbon foam was found to increase the corresponding maximum power density by 38.5 and 2.4 times than the uncoated and Pt C catalyst-coated electrode, respectively (Yu et al., 2015). Incorporation of catalyst such as Pd RGO CoFe2O4 nanoparticle, carbon-based RGO/MnO2, and Fe/Mn/C/F/O with PVDF on carbon fiber cloth was also found to be performed well in electricity production and minimizing membrane fouling in integrated MFC MBR system (Li et al., 2016; Gao et al., 2018a,b). An overflow type BES MBR system with membrane fouling mitigation by electricity produced by the MFC was developed by Zhou et al. (2015). It effectively showed that integration of MFC as pretreatment unit to MBR affects the sludge properties, such as decreasing particle zeta potential, particle size macroaggregation, extracellular polymeric substances (EPS) and SMP reduction, ratio of MLSSproteins/MLSScarbohydrates increase, which lead to membrane fouling mitigation (Zhou et al., 2015). Another interesting outcome of this investigation is a detailed illustration of the bacterial community involved in cathodic biofilm formation. Lactococcus sp. (28.3%) was predominant followed by Bacillus sp. (12.3%), Pseudomonas sp. (8.8%), Saprospiraceae_uncultured (8.4%), and Solibacillus sp. (6.8%) in the biofilm. The electroactive Gram-positive Lactococcus excretes membrane allied quinones to mediate electron transfer to extracellular electron acceptors, the presence of those can be expected due to continuous solution transport from anodic chamber to the cathodic chamber in this overflow type BES MBR (Freguia et al., 2009). The presence of Bacillus sp. shows that the system involved in simultaneous aerobic nitrification/denitrification (Joong et al., 2005) and Saprospiraceae is an important microorganism for protein degradation (Xia et al., 2008), which leads to higher NH41 N and total nitrogen removal efficiencies for this kind of systems. On the other hand, Pseudomonas is a well-documented aerobic bacterium present in MBR for degrading organic load from the system (De Gusseme et al., 2011). The membrane-cum-cathode module was also experimented to be attached with the MFC, treating the excess sludge generated from MBR, for facilitating the sludge

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reduction (Wang et al., 2018). The results showed 2.1 times higher power density, 50% lower internal resistance and 1.5 times increased sludge reduction rate, achieved by additionally introducing inclined anode in the settling tank along with 21.51% and 17.80% reduction in COD and SMPs in the supernatant, respectively. Thus membrane fouling was also mitigated significantly. In addition, packing of the membrane with stainless steel mesh, encompassed by granular activated carbon (GAC), further reduced the membrane fouling due to the negatively charged field around it (Wang et al., 2018). In another experiment by Katuri et al. a combined system of MEC with membrane filtration with an electrically conductive nickelbased hollow-fiber membrane for both the purposes of the cathode, and the filtration membrane was developed. The system could recover up to 71% of the substrate energy in the form of 83% methane-rich biogas, when applied with a voltage of 0.7 V. Authors also claimed that the biofouling in this system was mitigated by increasing rates of biogas production that can be controlled by regulating the applied voltage (Katuri et al., 2014).

12.4.1.2 The membrane as anode-cum-filtration unit The use of anode of an MFC as MF membrane (anode filter) in sidestream crossflow configuration in a tubular filtration module was developed by Madjarov et al. (2018). A commercial cross-flow filtration setup was fitted with a polypropylene filter cassette to accommodate the anode filter. Stainless steel filtration membrane operated with a pore size of 1 and 0.5 μm showed 11 and 15 A/m2 of current density, respectively, which was almost four times higher than the system without the presence of anode filter. The permeate flow was identified as the main parameter leading to increased current densities. The convective transport of protons away from the anode because of permeate flow prevented local drop of pH that can lead to lower current densities. The study thus showed that lower pH is not only thermodynamically unfavorable but also effectively inhibits the growth of exoelectrogens such as Geobacter sulfurreducens at pH values of 5.5 and below (Kim and Lee, 2010). Increased permeate flow further improved substrate supply at the bacteria electrode interface that was hypothesized to be another possible reason behind higher current density in the presence of the anode filter. Another interesting investigation by Li et al. (2014) showed that an electrically conductive low-cost polyester filter cloth, modified by in situ formed PANI PA, can be used as a cathode as well as anode material with bifunctional activity. Initial testing was done using graphite granule as the anode and the synthesized material as cathode, which gave a maximum power density of 44.80 mW/m2. When the graphite granule anode was replaced by a modified membrane, power density of 13.02 mW/m2 was obtained with inoculated Shewanella oneidensis. This PANI PA modified polyester filter cloth executed high conductivity, good filtration, and antifouling properties as electrode-cum-filtration membrane in MFC (Li et al., 2014).

An overview of membrane bioreactor coupled bioelectrochemical systems

e

259

e Ion-selective membrane

Influent

a

b

c

d Cathode

Anode Anaerobic compartment

e

Membrane filtration

Aerobic compartment Effluent

Figure 12.3 Schematic of basic functioning of hybrid BES MBR system through integration of membrane unit either as (a) anode-cum-filtration unit, (b) separator-cumfiltration unit, (c) cathode-cum-filtration unit, or (d) post-treatment unit to BES. BES, Bioelectrochemical system; MBR, membrane bioreactor.

12.4.1.3 The membrane as separator-cum-filtration unit The MBR incorporated with different electrode parts of BES in the integrated system has already been discussed; however, the use of membrane filtration unit as a separator to the BES can also be a viable option. The first of its kind research had been done by Wang et al. (2013c). They studied the use of integrated MFC MBR system without aeration for energy recovery and wastewater treatment using the nonwoven cloth as both the separator and membrane filter in the electrochemical MBR. Synthetic wastewater was continuously fed into the anodic chamber and treated water then flowed through the separator to the cathodic chamber. Fluid leakage through the cathode, pH gradient, and accumulation of inorganic salts deposits on the cathode, which are commonly encountered in air-cathode MFCs, could be overcome by such systems as claimed by the author. The system performed efficiently by removing 89.1% of COD and yielding a small amount of sludge and about 7.5% of the energy stored in the wastewater could be recovered as electrical energy (Wang et al., 2013c). The hybrid membrane module with bifunctional ability as separator and filtration membrane is also being extensively evaluated by the authors (Bhowmick and Ghangrekar, 2018) (Fig. 12.3).

12.4.2 Combined bioelectrochemical system membrane bioreactor system In the previous section a detailed discussion was done on the research work going on the replacement of a part of BES by an MBR unit in the integrated BES MBR system. Whereas, in this section, a detailed discussion will be made on the

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combination of these two processes as complementary to each other either as a pre- or posttreatment unit.

12.4.2.1 Membrane bioreactor as pretreatment unit The exploration of having MBR as a pretreatment unit to the BES has rarely been investigated. An AnMBR was developed by Dhar et al. (2013) followed by dualchamber MEC and operated with an HRT of 1 8 days, which showed an increment in current density values from 7.5 to 14 A/m2. Despite the fact that the membrane separation eradicated the presence of methanogens in the anodic chamber reducing the CH4 formation, fermenters for proper propionate utilization in MEC also got mitigated as per the researchers of this investigation. Fermenters break down the propionate into acetate and hydrogen gas, which were further utilized by anode-respiring bacteria. Hence, the reduced propionate fermentation rate limited the current generation in the MEC, which indicates the significance of balanced microbial structures in the anodic biofilm (Dhar et al., 2013). In another investigation by Borea et al., the activated sludge from an MBR at different TSS concentrations (1 10 g TSS per liter) was fed to run MFC and to assess the electrochemical response of the system. The results demonstrated that 30% higher COD removal was achieved as compared to the control system, which might be due to the influence of SS content to increase the current density to about 2.0 A/m2 (Borea et al., 2017).

12.4.2.2 Membrane bioreactor as post-treatment unit The most common form of combined MFC MBR system comes up with MBR as a post-treatment unit to BES, which facilitates the system by drastically reducing the membrane fouling phenomenon as discussed in this subsection. Wang et al. developed the first of its kind combined MFC MBR system, which favored a better utilization of the oxygen in the aeration tank of MBR by employing the low-cost MFC biocathode and enabled a high effluent quality. The aeration tank was used as the cathodic chamber with carbon felt as an electrode. Nylon mesh with a pore size of 74 μm was used as the filter media of MBR, which was placed at the bottom of the aeration tank. An average current of 1.9 6 0.4 mA was generated along with a maximum power density of 6.0 W/m3 by this system (Wang et al., 2012). In another investigation a modified MFC tubular MBR integrated system using the tubular membrane for cross-flow filtration as well as for simultaneous bioelectricity production and wastewater treatment was developed (Wang et al., 2013a). The tubular membrane module was placed outside the aeration tank. The system showed a maximum voltage output of 0.8 V and maximum power density of 0.04 W/m2 while removing 94% organic matter and more than 80% ammonia nitrogen from the wastewater (Wang et al., 2013a). Performance evaluation of two-stage MFC-anaerobic fluidized bed MBR was done by Ren et al. (2014) to produce high-quality effluent with minimal energy demands. Two MFCs were hydraulically connected in series, consisting of three anodes for each MFC made of graphite fiber brushes with a titanium wire core.

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The outlet of the second MFC was connected to anaerobic fluidized bed MBR. An average of 92.5% of COD removal was achieved at a combined HRT for 9 h, and nearly complete removal of TSS was demonstrated by this system. The energy required for anaerobic fluidized bed MBR was also less than aerobic MBR. At high flux rate and low HRT, in the laboratory-scaled operation, the energy produced by the MFCs (0.0197 kW h/m3) was claimed to be theoretically sufficient to meet the energy demand for the system operation itself (Ren et al., 2014). Kim et al. (2016) also examined the performance of anaerobic fluidized MBR to treat effluent of MFC. The GAC was used in this study as the fluidized particles in anaerobic fluidized MBR to support bacterial growth along with PVDF hollow-fiber membrane of 0.1 μm pore size. Authors claimed to achieve controlled membrane fouling by using fluidized GACs. The combined system further achieved 89% COD removal with effluent of 36 6 6 mg COD per liter over 112 days of operational period (Kim et al., 2016). The post-treatment of anodic effluent from MFC in MBR was also extensively investigated by the authors elsewhere (Bhowmick et al., 2019a,b) Based on the performance of membrane bioelectrochemical reactor using an anion exchange membrane (AEM) or cation exchange membrane (CEM) placed directly in the cathodic chamber for better nitrogen removal, Li and He (2015) claimed that by reducing the anolyte recirculation rate, the energy consumption by the system can be reduced. The AEM further showed 49.3% higher nitrogen removal than the CEM while maintaining a low membrane fouling due to the removal of organic compounds in the anodic chamber (Li and He, 2015). The organic matter could act as an electron donor in the anodic chamber, while denitrification in the cathodic chamber could be simultaneously achieved in a low-cost electrochemical MBR (Ma et al., 2015). The treated wastewater was obtained through MF membranes of 0.1 μm pore size, and overall HRT of the system adopted was 15.3 h. The nitrogen reduction organisms were notably enriched in the cathode biofilm of the electrochemical MBR due to the capture of electrons that were then used at the cathode to drive the denitrification process, which are important in autotrophic denitrification by which the system achieved 78.2% total nitrogen removal efficiency, despite the decrease in organic loading rate for heterotrophic denitrification (Ma et al., 2015). There are some exceptional cases, where the MBR is not exactly considered as pre/posttreatment unit rather than a combined intermediate one. Like in a study by Su et al. a combined system of MFC and MBR was developed, which showed a COD and ammonia removal efficiency of more than 90%, simultaneously achieving 5.1% higher sludge reduction, because of treatment of sludge in MFC, than that of the conventional MBR along with achieving an average voltage and maximum power density of 430 mV and 51 mW/m2, respectively. The sludge from MBR was fed to MFC and then again part of it was recycled back to MBR for proper sludge management. The MFC was fitted with carbon cloth electrodes with 10% platinum catalyst and coated with polytetrafluoroethylene as a gas diffusion layer. In this single-sludge MFC MBR, about 75 mg/L COD could be translated to electricity in one cycle, and membrane fouling was mitigated by the sludge modification in this combined MFC MBR system (Su et al., 2013). In another investigation, the hollow-fiber UF membranes were placed inside the anodic chamber of tubular MFC (Ge et al., 2013).

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The system achieved 90% COD removal and produced effluent with less than 1 NTU turbidity while treating synthetic wastewater producing 0.003 0.025 kW h/m3 normalized energy. The PVDF hollow-fiber membranes of 0.02 μm were placed around the carbon brush electrode and operated intermittently to reduce the membrane fouling. The performance details of various BES-MBR systems investigated so far by various researchers are presented in Table 12.1.

12.5

Outlook and future perspectives

The investigations on the application of BES MBR systems done so far in the field of wastewater treatment demonstrated higher efficacy of the system by complementing the benefits of each component of it. Though the treatment regime is improved with enhanced net energy output, there are several issues/challenges that need to be addressed for making this technology a commercial success.

12.5.1 Water-energy nexus The relation between the wastewater treatment and energy production are complexly interconnected, which is not exactly a closed loop as the amount of water used for different energy conversion processes need not always be the same. The BES MBR system can be a viable replacement to the energy-consuming existing wastewater treatment systems with more neutral to positive water-energy nexus, only when the net energy recovery is properly evaluated and the further investigation should proceed accordingly to minimize the energy need or to make it net energy neutral process.

12.5.2 Membrane fouling mitigation The main challenge, all the filtration technologies encounter, is the biological or chemical fouling of membranes used. Although most of the investigations in this BES MBR advocated the mitigation in membrane fouling, the investigation should go in finding proper relation between fouling and electricity produced by the system. Like how bioelectricity affects membrane fouling and vice versa. Most of the study showed that the bioelectricity produced by the BES can reduce the membrane fouling by surfacing it with electrical charge field to restrain oppositely charged microorganisms or by degrading organic molecules, or by lowering anodic pH, or by producing reactive oxygen species, which may inhibit the formation of biofouling on the membrane surface.

12.5.3 Control of emerging contaminants The emerging contaminants are defined as chemicals that are not currently (or have been only recently) regulated by the environmental regulatory bodies and about which there are concerns regarding their impact on human or ecological health.

Table 12.1 Performance evaluation of different bioelectrochemical system (BES) membrane bioreactor (MBR) systems. Arrangement/ configuration

Membrane type

Pore size

Wastewater

Power density

Removal efficiency COD (%)

Nitrogen (%) 95.6 of NH4 N

Remark

References

40 μm

Synthetic

4.35 W/m3

92.4

0.02 μm

Synthetic

, 2 W/m3

90

Ge et al. (2013)

Filtration media

74 μm

Synthetic

6.0 W/m3

89.6

Ultrafiltration membrane

Filtration media

0.04 μm

Synthetic

Wang et al. (2012) Dhar et al. (2013)

Platinum-based multiwalled carbon nanotubes ultrafiltration membrane Hollow-fiber microfiltration membrane

Cathode and filtration media

40 μm

Domestic wastewater

6.8 W/m3

Filtration media

0.1 μm

Synthetic

51 W/m2

. 90

. 90 of NH4 N

Separator and filtration media Cathode and filtration media

50 μm

Synthetic

7.4 W/m3

89.1

51.3

48 μm

Synthetic

8.62 W/m3

86.1

97.5 of NH4 N

Bioelectrochemical MBR

Stainless steel mesh

Membrane bioelectrochemical reactor MFC MBR

PVDF hollow-fiber membranes Nylon mesh

Anaerobic MBR as a pretreatment to MEC Air-biocathode MFC

Sludge MFC stack and MBR

Membrane role

Electrochemical MBR

Nonwoven cloth

Bioelectrochemically assisted MFC

Stainless steel membrane

Cathode and filtration media Filtration media

Wang et al. (2011)

53

97 SCOD

97 of NH3N

Malaeb et al. (2013)

Higher sludge reduction than conventional MBR

Su et al. (2013)

Wang et al. (2013c) Effluent turbidity of 0.8 NTU

Wang et al. (2013d)

(Continued)

Table 12.1 (Continued) Arrangement/ configuration

Membrane type

Membrane role

Pore size

Wastewater

Power density

Removal efficiency COD (%)

Nitrogen (%)

MFC and tubular MBR

PVDF tubular membrane

Filtration media

0.1 μm

Synthetic

0.040 W/m2

94

80 of NH4 N

Tubular electrochemical MBR

Stainless steel mesh

Cathode and filtration media

40 μm

Synthetic

1.43 W/m3

93.7

96.5 of NH4 N

Anaerobic electrochemical MBR

Nickel-based hollowfiber membrane

Cathode and filtration media

1 μm

Synthetic

MFC MBR

Polyester filter cloth, modified by phytic acid Stainless steel mesh

MFC with independent membrane cathode bioreactor MFC and anaerobic fluidized bed MBR Anaerobic membrane bioelectrochemical reactor MBR MFC

PVDF hollow-fiber membranes Hollow-fiber microfiltration membranes Carbon fiber cloth

Cathode and filtration media Filtration media Cathode and filtration media

. 95

Remark

H2O2 produced at cathode contributed in fouling mitigation by removing the membrane foulants 0.27 kW h/m3 of energy was required for this system, which is much less than aerobic MBR

References

Wang et al. (2013a) Wang et al. (2013b)

Katuri et al. (2014)

Synthetic

784.08 mW/ m3

95

Li et al. (2014)

15 μm

Synthetic

0.15 W/m3

. 90

0.1 μm

Domestic wastewater

0.0197 kW h/ m3

92.5

0.4 μm

Synthetic

1.16 W/m3

91.6

94.8 of Nitrate

Reduced membrane fouling

Tian et al. (2014)

5 20 μm

Synthetic

1358 mW/m3

90

80 of NH4 N

Cathode filter with a catalyst containing C, Mn, Fe, and O

Li et al. (2015b)

Reduced membrane fouling

Liu et al. (2014)

Ren et al. (2014)

MFC MBR

MBR with ionexchange membrane

Electrochemical MBR

MFC and intermittently aerated biological filter MFC with hollowfiber MBR

MFC-aerobic MBR coupled system

Hybrid MBR MFC

Hollow-fiber ultrafiltration membrane Anion-exchange membrane

Filtration media

Synthetic

Filtration media

Synthetic

Microfiltration membrane

Filtration media

0.1 μm

Filtration media

5.49 W/m3

Li et al. (2015a) 91.3

56.9

78.2

Municipal wastewater

98.4 mW/m2

80.1

Synthetic

0.27 kW h/ m3

91.7

Filtration media

0.03 μm

Synthetic

2.18 W/m3

. 90

. 90

Polyester filter cloth modified with the conducting anthraquinone disulfonic salt Polyester filter with carbon foam Fe Co

Cathode and filtration media

9.8 μm

Synthetic

0.35 W/m3

92.5

70.6 of NH4 N

Cathode and filtration media Filtration media

1 μm

Synthetic

135 mW/m2

95

85 of NH4 N

80 nm

Synthetic

1.02 W/m3

98

Cathode and filtration media

38 μm

Synthetic

629 mW/m3

92.6

Hollow-fiber polysulfone

Overflow-type electrochemical MBR

Stainless steel mesh

Li and He (2015)

Ma et al. (2015)

Dong et al. (2015)

PVDF hollow-fiber membrane

MFC MBR

Anion exchange membrane showed better nitrogen removal than cation-exchange membrane Enrichment of nitrogen reducing organisms in cathode biofilm

73.9

MFC MBR improved dewaterability and filterability of the sludge Reduced membrane fouling due to MFC integration and H2O2 production Low-cost electrode base material and catalyst

Reduced membrane fouling due to MFC integration

Tian et al. (2014)

Xu et al. (2015)

Yu et al. (2015)

Ghosh Ray et al. (2016) Zhou et al. (2015)

(Continued)

Table 12.1 (Continued) Arrangement/ configuration

Membrane type

Membrane role

Pore size

Wastewater

Power density

Removal efficiency COD (%)

UASB coupled with hybrid aerobic MBR

A hollow-fiber ultrafiltration membrane

Filtration media

0.04 μm

Synthetic

95

MFC followed by anaerobic fluidized MBRs

PVDF hollow-fiber membrane

Filtration media

0.1 μm

Domestic wastewater

89

Integrated MFC MBR

Carbon fiber based Pd RGO CoFe2O4 catalyst

50 120 nm

Synthetic

506 mW/m3

. 95

Hybrid MFC MBR

PVDF hollow-fiber membrane

Cathode and filtration media Filtration media

0.22 μm

Synthetic

0.059 W/m2

53.71

Electrochemical process with MBR

Hollow-fiber ultrafiltration module

Filtration media

0.04 μm

Synthetic

Fe/Mn/C/F/O catalytic cathode in BES

Carbon fiber based PVDF with Fe/Mn/ C/F/O membrane

2 7 nm

Synthetic

MnO2 catalyst cathode

Carbon-based RGO/ PVDF/MnO2 membrane

Cathode and filtration media Cathode and filtration media

30 35 nm

Synthetic

Remark

References

Higher organic micropollutants removal in anaerobic condition GAC in anaerobic fluidized MBRs to support bacterial growth

Alvarino et al. (2016)

Nitrogen (%)

. 85 of NH4 N

98

72.10 of NH4 N

446 mW/m3

. 97.4

. 96.7 of NH4 N

228 mW/m3

. 97.4

. 97 of NH4 N

Kim et al. (2016)

Li et al. (2016)

0.5 mA cm2 of external power applied, removal of pharmaceuticals Quartz sand separator between anodic and cathodic chamber Quartz sand separator between anodic and cathodic chamber

COD, Chemical oxygen demand; GAC, granular activated carbon; MEC, microbial electrolysis cell; MFC, microbial fuel cell; PVDF, polyvinylidene fluoride; UASB, upflow anaerobic sludge blanket; SCOD, Soluble COD.

Wang et al. (2016) Ensano et al. (2016)

Gao et al. (2018a)

Gao et al. (2018a)

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A few examples of emerging contaminants include disinfection by-products, pharmaceutical, and personal care products, persistent organic chemicals, and mercury as well as their degradation products. Feng et al. (2014) demonstrated that the removal of p-fluoronitrobenzene was significantly enhanced by the cathodic reduction in a BES. In another study, Ensano et al. (2019) explored that the removal efficiency of humic substances, NH4 N, and orthophosphate in the BES MBR was 90.7%, 72.1%, and 100%, respectively, along with a reduction in the content of diclofenac, carbamazepine, and amoxicillin by 75.3%, 73.8%, and 72.1%, respectively. Hence, it is expected that BES MBR system has great promise for removing emerging contaminants that can be further explored in detail.

12.5.4 Field-scale applications The practical applications of BES MBR systems are subject to scaling up that remains a significant challenge, because of its configuration of both integrated or separately attached membrane modules and low energy output from BES. Proper kinetic modeling with detailed life cycle analysis is required to be done to help optimize the reactor configuration and predict system performance for any possible field-scale applications.

12.6

Conclusion

Bioelectrochemical treatment technology has a wide range of applications in the wastewater treatment field in a sustainable way. The performance of BES is further enhanced by combining MBR with it to obtain higher quality effluent and to reduce the energy demand along with lessening the footprint of the system. The use of MBR along with BES has a higher potential in removing suspended solids and organic matter as compared to a single BES, which makes it a reliable and viable treatment option prior to reuse or recycle. The hybrid system of BES MBR has also shown better performance for removal of nutrients, pathogens, and antibiotics from the wastewater. The MFC, on the other hand, had effectively complemented MBR by stabilizing the sludge and controlling the fouling of the membrane. Also, the membrane filtration unit is flexible to be incorporated in the cathodic or anodic chamber, or it can be used as a separator between the electrodes or as an external arrangement in the BES as per the requirement. The MBR alone is an energyintensive process; however, the hybrid system of BES MBR can make it a selfsustainable system for a viable alternative to the existing wastewater treatment systems in the near future for field-scale applications.

Acknowledgment We gratefully acknowledge the Department of Biotechnology, Government of India for providing financial support under the INNO INDIGO scheme (BT/IN/INNO-INDIGO/28/MMG/ 2015-16).

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Integration of microbial fuel cell into constructed wetlands: effects, applications, and future outlook

13

Supriya Gupta1,3, Pratiksha Srivastava2 and Asheesh Kumar Yadav1,3 1 Academy of Scientific and Innovative Research (AcSIR), CSIR-Human Resource Development Centre, (CSIR-HRDC) Campus, Ghaziabad, India, 2Environment and Sustainability Department, Australian Maritime College (AMC), College of Science and Engineering, University of Tasmania, Launceston, TAS, Australia, 3CSIR-Institute of Minerals and Materials Technology, Bhubaneswar, India

Chapter Outline 13.1 Introduction

273

13.1.1 Probable electron transfer mechanism in constructed wetlands-microbial fuel cell 275 13.1.2 Basic characteristic of constructed wetlands and their similarity with microbial fuel cell 276

13.2 Development of merger technology

278

13.2.1 Design and operation of constructed wetland-microbial fuel cells 279 13.2.2 Performance assessment of constructed wetland-microbial fuel cells 281

13.3 Challenges and future perspectives Acknowledgment 288 References 288

13.1

287

Introduction

The treatment of wastewater has always been a cost-intensive process, irrespective of its source. The continuous growth in the population, industries, and urban development is further increasing the demand for potable and ultimately leads to the production of an increased amount of wastewater, which is finally threatening the environment. The constructed wetland (CW) in such a situation stands as a sustainable and economical wastewater treatment technology. CWs are the human-made or engineered system with the possibility of operational controllability. CWs “no chemical” and “no electricity” requirement feature make them low-cost technology. Moreover, CWs are easy in operation and maintenance, which adds up to their advantages as a decentralized treatment system for small communities and households (Garcı´a et al., 2001). Thus amidst the water
energy crisis, CWs are becoming a suitable alternative to conventional wastewater treatment technologies and are being used throughout the world (Wallace and Knight, 2006; Garcia et al., 2014). Since their identification as a wastewater treatment system, CWs have been differently designed and timely modified for the treatment of a large range of Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00013-8 © 2020 Elsevier Inc. All rights reserved.

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wastewaters, including domestic wastewater, industrial wastewater, urban wastewater, agricultural wastewater, stormwater runoff, leachates, and mine drainage (Vymazal, 2014a,b). In the 1950s Seidel started her work with free water surface (FWS) wetlands using wetland plants for the treatment of different types of wastewater. Later, in the 1960s, she intensified her trials on wetlands and initiated the use of horizontal subsurface flow (HSSF) CWs (Vymazal, 2005). In the early phase, FWS CWs dominated in North America, whereas HSSF CWs in Europe and Australia (Vymazal, 2005; Brix, 1994; Vymazal, 2011). Some of the works of Seidel gave the idea of hybrid systems, and at the end of the 20th century, hybrids of horizontal and vertical flow SSF CW came into the picture (Vymazal, 2005). The initial studies on various types of CW had revealed the importance of the presence of oxygen (as an electron acceptor) for the enhancement of treatment efficiency and, thus, the role of macrophytes in it. With the outlook of fastening the treatment performance of CWs, in less complicated and/or probably economically viable system, there have been many efforts for the improvement of treatment efficiency of CWs that can be categorized as (1) operation strategies such as effluent recirculation, artificial aeration, tidal operation, drop aeration, flow direction reciprocation, earthworm integration, bioaugmentation; (2) configuration innovations such as circular-flow towery hybrid CW, baffled subsurface-flow; and (3) electron donor supplementation (Wu et al., 2014). These efforts improved the treatment performance of CWs. However, certain limitations still remained. For instance, aeration (gas bubbling) resulted in excessive microbial activities and growth, flushing out of suspended solids, clogging, etc. (Chazarenc et al., 2009). Moreover, aeration itself is a costintensive process. In general, the CWs are dominated by anaerobic pollutant removal mechanisms that are slow as compared to the aerobic reactions. This slowness of anaerobic reactions is probably due to the absence of an adequate availability electron acceptors or suitable electron acceptors in CW (Srivastava et al., 2015). The implementation of a suitable temporary electron acceptor in the inner anaerobic depth of CW can thus enhance its treatment efficiency. In the same context the recently emerged CWs integrated with microbial fuel cells (MFCs) (Srivastava et al., 2015, 2019; Yadav et al., 2012) have been proven as an efficient technology for pollutant removal along with electricity generation in the treatment wetlands. In said concept the anode of MFC acts as the electron acceptor in the deep anaerobic region of CW and speeds up the slow anaerobic processes. The enhancement is a result of coupling the oxidation and reduction reactions in anaerobic and aerobic regimes, respectively, which positively affects the kinetics of the microbial processes taking place. The following equations are showing the oxidation of glucose at anode and reduction of oxygen at the cathode (Logan and Rabaey, 2012; Wang and Ren, 2013). Anode: C6 H12 O6 1 6H2 O ! 6CO2 1 24e2 1 24H1

(13.1)

Cathode: O2 1 4e2 1 4H1 ! 2H2 O

(13.2)

Furthermore, the enhanced reaction in CW-MFC would decrease the large land footprint requirement.

Integration of microbial fuel cell into constructed wetlands: effects, applications, and future outlook 275

The segregated redox condition in anode and cathode chambers of MFC is the basis for electric generation (Hooker, 1933). The primitive MFCs were merely electricity-generating devices that were fueled with chemicals. Gradually, the potential of wastewater as a resource was realized (Abbassi and Abbassi, 2012; Tauseef et al., 2013), and the fueling of MFCs drifted from chemicals to organic matter enriched wastewater. The establishment of an anode (electron acceptor) in the lower bed and cathode (electron donor) in the upper bed of CW, connected with electric wire, directs the flow of microbial generated metabolic electrons in the anaerobic region to the aerobic region. The presence of anode, coupled with the cathode in CW, helps in increasing the respiration. Thus with the integration of MFC into CW, the rate-limiting anaerobic reactions in the inner bed of CWs get enhanced by using electroconductive anode as an artificial electron acceptor (Ramı´rez-Vargas et al., 2018; Srivastava et al., 2018b,c). In recent years, CW-MFC has emerged as an intensified wastewater treatment technology. The responsibility of efficient electron transfer in conductive material
based treatment was found to be dominated by electrogens—a group of specialized microbes for electron transfer (Lovely, 2006, 2008; Nevin et al., 2010). The detailed electron transfer mechanism is described in the subsection.

13.1.1 Probable electron transfer mechanism in constructed wetlands-microbial fuel cell The generation of electrons mainly occurs in the anaerobic chamber, where the chemical bonds of organic matter get oxidized by electrogens and liberate electron and proton. The electrogens, also known as “electrigens” or “electroactive bacteria” (EAB) or “exoelectrogens,” are specialized microbes that can respire electrons extracellularly, generated from the oxidation of organics. Furthermore, the microbial action is basically based on the interactions with electron conductive materials, that is, insoluble electron acceptors and donors. Electrogens dump the generated electron to the outside of the cell on an electron acceptor such as conductive material, either via direct or indirect transfer mechanisms (Logan, 2009). Fig. 13.1 shows the detail about the electron transfer mechanism in electrogens. Direct electron transfer most likely takes place either by outer-membrane redox proteins and cytochrome cascades or by conductive nanowires (Bonanni et al., 2012; Busalmen et al., 2008), whereas indirect electron transfer occurs by metabolically generated electron shuttle or mediator by electrogens that are later used for extracellular electron transfer (Arends and Verstraete, 2012). These mediators/shuttles molecules can be either generated by bacteria themselves, such as humic acid and fulvic acid, or provided artificially such as 2-amino-3-carboxy-1,4-naphthoquinone, anthraquinone-2,6-disulfonate (AQDS) (Mao and Verwoerd, 2013; Watanabe et al., 2009). These shuttles have been studied extensively to find their role in bioremediation of metals, halogenated organics, azo dyes, etc. and enhancement of electricity generation in MFCs. It has been found that natural mediators, humic acid and fulvic acid, promoted the oxidative degradation of organic compounds such as 4-chlorophenol (Kang and Choi, 2009), methyl tert-butyl ether, and tert-butyl

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Figure 13.1 Electron transfer mechanisms in exoelectrogens: direct electron transfer through direct contact of microbe with the electron acceptor (A and B), and indirect electron transfer using self-generated or artificial mediators (C).

alcohol (Finneran and Lovley, 2001). In addition, these have also promoted the reduction of hexahydro-1,3,5-trinitro1,3,5-triazine (RDX), a widely used explosive (Kwon and Finneran, 2008) and azo dyes such as Remazol Black B dye (Pearce et al., 2006). Artificial mediator AQDS has also been studied for the enhanced degradation of RDX and Remazol Black B dye (Kwon and Finneran, 2008; Pearce et al., 2006), and reduction of metal (Luo and Gu, 2009). Similarly, the role of shuttles in enhancing the current generation in MFCs has also been demonstrated by Reguera et al. (2006) and Gorby et al. (2006). The two well-known and studied bacterial genera of exoelectrogens are Shewanella (Ringeisen et al., 2006) and Geobacter (Kiely et al., 2011), which are found to be electrically active in all types of conductive material
based wastewater treatment.

13.1.2 Basic characteristic of constructed wetlands and their similarity with microbial fuel cell The understanding of biochemistry is very important to know the complexity of microbial action, the degradation, and the electron transfer mechanism of

Integration of microbial fuel cell into constructed wetlands: effects, applications, and future outlook 277

CW-MFCs. CW, in general, is the simply lined pit assembled with water-loving plants (macrophytes) and stones/gravels with microbial biofilm surfacing them, along with water flowing through (Corbella and Puigagut, 2016; Vymazal, 2005). Fig. 13.2 gives a clear picture of CW. These components of CW are crucial for making it proficient in the removal of many contaminants such as organics, nitrogen, phosphate, heavy metals, and pathogens from the wastewater (Vymazal et al., 1998; Sakadevan and Bavor, 1998). Furthermore, the vegetation and filter medium (primary biological components of CW) provide anchorage to the thriving microbes to form a fixed biofilm over them (Gorgoglione and Torretta, 2018). Similarly, plant roots (rhizomes) present in the upper regime of CWs transfer oxygen from the atmosphere to the inside of the filter bed. In addition, the plant roots create a rich network of aerobic micro-sites in its vicinity and thereby retain aerobic conditions in the upper regime (Gorgoglione and Torretta, 2018). The atmospheric oxygen also dissolved in the water by diffusion and convection in the CW to a certain extent (Brix, 1993). This generates the oxygen concentration gradient and thus redox gradient in the upper and the lower regimes of the filter bed of CWs, naturally. The wastewater flows through CWs is subjected to various biophysiochemical processes, such as sedimentation, filtration, precipitation, coagulation, and plant uptake, and various microbial processes such as adsorption and volatilization (Vymazal, 2007; Kadlec and Wallace, 2009; Faulwetter et al., 2009; US EPA, 2000). However, the redox potential (Eh) conditions, along with different loading rates, temperatures, soil types, and operation strategies, directly and indirectly, influence these processes (Biederman et al., 2002; Stein et al., 2003; Stein and Hook, 2005; Yang et al., 2011). CW is further categorized based on the flow regime, namely, SF and SSF. The SSF-CW is more popular than SF due to its robustness. The Eh of SSF-CWs varies extremely, spatially, and temporally,

Figure 13.2 Cross section of HF-SSF CW picturizing the various components and the anchorage of microbial biofilm over the plant rhizome and gravels. CW, Constructed wetland; SSF, subsurface flow; HF, horizontal flow.

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additionally, it is influenced by several other factors such as the presence of plants, fluctuations in the water level due to evapotranspiration, light intensity, and temperature (Wiessner et al., 2005; Duˇsek et al., 2008; Bialowiec et al., 2012; Garcia et al., 2010). However, the Eh variability in SSF-CW is high (fluctuation of several hundreds of millivolts within few hours) (Duˇsek et al., 2008) in-depth rather than across the length of the bed (Garcı´a et al., 2003). Since the Eh determines the pollutant removal, SSF CWs can be designed, by changing operational and design parameters, to favor the desired range of redox conditions for specific pollutants (Faulwetter et al., 2009; Gorgoglione and Torretta, 2018). Furthermore, recently, the controllability of Eh is also found to be interesting for electricity generation in CW-MFCs (Doherty et al., 2015a; Yadav et al., 2012). In this regard the microbial action is very much dependent on the potential of anode and cathode zones. Interestingly, the CW itself is having the capability to maintain the redox potential that favors the intensified pollutant removal and electricity generation from CW-MFCs (Corbella et al., 2014). Furthermore, the chemistry of electrogens relies on the exchange of metabolically generated electrons that are removed from an electron donor and supplied to an electrode acceptor (e.g., a solid-state electrode) through an electroconductive material. This flow of electron between the donor and acceptor is owed to the difference in redox potential, that is, toward a high redox electron acceptor, such as oxygen (Corbella et al., 2014). Therefore the CW-MFC merger technology fuses microbiology and electrochemistry, where interactions between living microbes and electrodes happen through capacitive/Faraday interactions (Ramı´rez-Vargas et al., 2018).

13.2

Development of merger technology

The demonstration of a new merger technology, CW-MFC, was for the first time done by Yadav (2010). The first detailed study on CW-MFCs was demonstrated by Yadav et al. (2012) in a laboratory-scale vertical CW for dye removal. The study has achieved 80.8% of dye removal and 75% of chemical oxygen demand (COD) removal at 1500 mg/L of initial azo dye concentration. The motivation was spurred from the similarity and compatibility of the two technologies in terms of substrate utilization, design configuration, and microbial activities involved. Later, Zhao et al. (2013) had investigated the performance of CW-MFCs based on continuous and batch flow and found the system that was running in continuous mode has achieved higher COD removal than the batch mode system. In the subsequent year, Villasenor et al. (2013) has performed a laboratory-scale HSSF CW-MFC and achieved 90%
95% of COD removal. Many other works of literature clearly state that the redox gradient is the overlapping feature of the two technologies (Corbella and Puigagut, 2016; Wang et al., 2017b). In MFC the membrane separates the two chambers and creates the gradient between them, whereas, in CW, the redox gradient exists naturally (of the order approximately 0.5 V) between the upper (air
water interface) and lower (completely/more anaerobic) regions of the bed.

Integration of microbial fuel cell into constructed wetlands: effects, applications, and future outlook 279

Figure 13.3 Schematic representation depicting the similarities in MFC and CWs configuration; and the possibility of implanting MFC inside CW as CW-MFC technology. CW, Constructed wetlands; MFC, microbial fuel cell.

Thus there is an opportunity for the strategic in situ implantation of the components of MFCs (anode as the electron acceptor and cathode as the electron donor, along with external load) within the CW matrix (Fig. 13.3). Like traditional MFC, in CW-MFC, the anode (electron acceptor) is placed at the anaerobic region and the cathode (electron donor) at the air
water interface of the oxic or aerobic region. The presence of electron acceptor in anaerobic region enhances the metabolic rate of EAB and promotes their growth. Thus the rate of organics degradation (electron donors such as glucose) increases and generates more electrons, protons, and carbon-dioxide molecules. The electrons generated are donated/transferred to the surface of anode electrode by EAB, and from there, these further transferred to the cathode electrode for the reduction of oxidants such as NO32 and O2 at the cathode, through an electroconductive wire. The proton generated during the oxidation of organics moves toward the cathode along with the hydraulic flow, where they combine with electron and O2 molecule to form H2O2 and H2O molecules. The formation of reduced products at the cathode, such as H2O2, can serve the purpose of disinfection and improves the quality of treated water further (Xu et al., 2016b).

13.2.1 Design and operation of constructed wetland-microbial fuel cells The design strategy of CW-MFCs is an important factor for wastewater treatment and electricity generation. The designing of CW-MFCs should be in the way to harvest more electrons from the oxidation of pollutants, provide proper interaction between the pollutants and electrode, avail the higher redox potential, and further the availability of terminal electron acceptor at the cathode. Many researchers have

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considered that an important parameter for designing CW-MFCs is flow regime. They have investigated CW-MFCs in HSSF (Villasenor et al., 2013), upflow (Doherty et al., 2015b,c), and mixed flow (Zhao et al., 2013). Villasenor et al. (2013) reported that the HSSF performed well with a low organic loading rate (13.9 g COD/m2/day). However, with the increase in loading rate, the cathodic region of CW-MFC suffered from an oxygen depletion state, and a sharp decline in a current generation was observed, although the pollutant treatment efficiency remained 80% or more. Subsequently, Doherty et al. (2015b,c) reported that vertical upflow hydraulic regime removed 79%
81% COD from wastewater (initial COD concentration of 0.411
0.854 mg/L) and achieved the power density of 0.168 W/ m3 (anode volume). Furthermore, they reported that simultaneous upflow
downflow CW-MFC in similar operational conditions removed COD only by 64% 6 4.6% but increased ammonium removal efficiency by 16%
20%. Also, the power density increased to 0.276 W/m3 (anode volume), namely, by 70% and reduced the internal resistance by 67% (500 to 300 Ω). Zhao et al. (2013) investigated the effect of mixed flow by providing aeration in the cathode along with maintaining the upward flow of wastewater and reported 76.5% COD removal (average influent COD concentration of 1058 6 420.89 mg/L) with a peak power density of 9.4 mW/m2. Inspired by the inherent designs of MFC and CW, the aforesaid configurations are set up as single-chambered with interelectrode separation of various materials. The vertical upward flow is the most commonly used design in which the anode constitutes the lower section, and the cathode is the upper region near the surface and plant roots (Guadarrama-Pe´rez et al., 2019). The deeply buried anode and superficially placed cathode at the air
water interface are separated by a large distance (membrane less), and sometimes by some separators such as glass wool (GW) and fiberglass material are also helpful in maintaining redox gradient for the functioning of CW-MFC. The use of GW has been done by Yadav et al. (2012), Zhao et al. (2013), and Doherty et al. (2015b,c) in CW-MFCs as it provides sharp redox gradient between the anode and cathode, with decreased interelectrode distance. The increased interelectrode distance in CW-MFC is responsible for higher ohmic resistance and decreased coulombic efficiency in comparison to conventional MFCs (Doherty et al., 2015b,c; Ahn and Logan, 2012; Srivastava et al., 2015). Villasenor et al. (2013) reported an internal resistance of 120 Ω using graphite plate electrodes with an interelectrode distance of 85.0 cm in a HSSF CW-MFC planted with Phragmites australis. Whereas Liu et al. (2005) reported the value of 35 Ω in conventional MFC, volume 28.0 mL, with the electrode spacing of 2.0 cm. Moreover, this value was decreased to 16.0 Ω when the electrode spacing decreased to 1.0 cm (Liu et al., 2005), though the oxygen diffusion toward the anode happened to result in decreased power output and coulombic efficiency. Doherty et al. reported a similar finding while studying the influence of the presence and absence of GW as a separator in CW-MFC. The higher voltage was achieved in the nonseparator system (465.7 6 4.2 mV with an electrode spacing of 5.0 cm), which was 48.9% higher than the highest value generated in GW separator
based system (312 6 7.0 mV with an electrode spacing of 2.0 cm). Also, the highest power density was produced in the nonseparator system (66.22 mW/m2), which was 3.9 times

Integration of microbial fuel cell into constructed wetlands: effects, applications, and future outlook 281

higher than the value in the GW system (17.14 mW/m2). It is probably because of the dense biofilm formation at the cathode that hindered the diffusion of oxygen from the open air. Thus the optimized distance between the anode and cathode, additionally, the flow of wastewater stream from bottom to top, maintains the requisite redox gradient (Corbella and Puigagut, 2016; Fang et al., 2013; Yadav et al., 2012) for maximum treatment and electric generation efficiency. For the design of a CW-MFC system to be profitable, several factors must be considered, such as the long-term operation, ease of management, and environmentally safe (Nitisoravut and Regmi, 2017; Sophia and Sreeja, 2017). Also, to enhance the performance of CW-MFC, few elements must be carefully addressed, such as substrate, type of membrane, external and internal resistance, electrode materials, and the space between electrodes. Dordio and Carvalho (2013) investigated the “good fit” for graphite granules and granular activated charcoal filler material in the CW matrix. It was found that the granule size has a major impact on current production since the particle size between 0.25 and 0.5 mm produced the current density 77.7 mA/m2, whereas the particle size of 1
5 mm produced only 37.9 mA/m2. Xu et al. (2016a) studied the influence of powdered activated charcoal (PAC)
modified dewatered alum sludge, in the anode. It was found that the 10.0% PAC addition in the alum sludge increased the COD removal by 10.0%, as compared to “no PAC system.” The CW-MFCs operated with intermittent aeration (IA)
supplemented cathode and radial oxygen loss (ROL)
assisted cathode were examined, and it was found that CW-MFC with IA achieved 78.71% and 53.23%, and CW-MFC dependent on ROL achieved 72.17% and 46.77% COD removal from synthetic wastewater containing glucose loads of 0.7 and 2.0 g/L, respectively. The maximum power density obtained was 31.04 and 19.60 mW/m3 in CW-MFC with additional IA and CW-MFC dependent on ROL, respectively (Srivastava et al., 2017). The different CW-MFC designs and configurations studied are presented in Table 13.1.

13.2.2 Performance assessment of constructed wetlandmicrobial fuel cells The synergy of MFC with CW not only retains its characteristic of upscaled wastewater treatment with low operation and maintenance investment but also enhances the treatment performance, as compared to traditional CWs, by linking its segregated regions and channelizing the flow of electrons that can be further harvested as electricity. Table 13.2 covers the studies focusing on the two aspects of this synergistic technology, that is, enhancing electricity generation and improving wastewater treatment. The studies by Srivastava et al. (2015) and Fang et al. (2013) support the concept. Srivastava et al. (2015) have also demonstrated that the operation of CW-MFCs in a closed circuit can enhance 12%
20% COD removal in comparison to the open-circuit operation. The performance enhancement of 12.6% was found for treating recalcitrant azo dye, ABRX3 wastewater (with 180 mg/L COD) (Fang et al., 2013). Doherty et al. (2015c), in his study, revealed the impact

Table 13.1 The detailed description of different CW-MFC designs and configurations studied, along with their treatment and power output efficiencies. Sr. no.

Reactor design and configuration

Wastewater type

Scale and HRT

Output descriptions

References

1.

Simultaneous upflow
downflow vertical CW-MFC Electrodes: graphite granule embedded graphite rod

Swine

8.1 6 0.12 L; HRT 5 1 day

Reduced the internal resistance by 67% to 300 Ω

Doherty et al. (2015c)

CW-MFC using PAC-modified alum sludge in anode chamber

Swine

1.5 L; HRT 5 1 day

2.

Anode: stainless steel mesh Cathode: graphite granule embedded stainless steel mesh

3.

An electroconductive horizontal subsurface flow biofilter acting as whole electrode Electrode: coke

Pretreated municipal

11 L; HRT 5 0.5
4 days

4.

Vertical CW-MFC IA and ROLassisted cathode (fed-batch mode)

Synthetic

1.8 L; HRT 5 1 day

Anode: graphite felt Cathode: Pt-coated carbon cloth

Increased the maximum power density by 70% to 0.268 W/m3 Increased ammonium removal efficiency by 16%
20% PAC in anode 5 0% COD removal 5 70% TN removal 5 32.5% MPD 5 41.39 mW/m2 Cathodic TEPS 5 44.59 μg/g wet sludge PAC in anode 5 10% COD removal 5 81.4% TN removal 5 44.9% MPD 5 87.79 mW/m2 Cathodic TEPS 5 87.7 μg/g wet sludge OLR 5 2.0
12.7 g BOD5/m2/day COD 5 93% BOD 5 99% NH4 5 97% TN 5 69% COD removal: 78.71% and 53.23% with 700
2000 mg/L organic load and (CW-MFC with IA); 72.17% and 46.77% (CW-MFC with ROL) Power density: 31.04 mW/m3 (anode volume) (CWMFC with IA); 19.60 mW/m3 (anode volume) (CW-MFC with ROL)

Xu et al. (2016a)

AguirreSierra et al. (2016) Srivastava et al. (2017)

Optimal interelectrode distance 5 20 cm with, COD removal 5 94.90% Power density 5 0.15 W/m3 Internal resistance 5 339.80 Ω Coulombic efficiency 5 0.31% In addition, optimal COD 5 200 mg/L with COD removal 5 90.45% OCV 5 741 mV Power density 5 0.20 W/m3 Internal resistance 5 339.80 Ω The COD removal efficiency was improved from 83.2% to 88.7%; and total nitrogen removal efficiency increased from 53.1% to 75.4%, as compared with single-stage CW-MFC

Song et al. (2017)

HRT 5 3 days

Avoid the immoderate elongation of emerged plants’ roots which would ruin electrode materials

Shen et al. (2018)

Synthetic

2 L; HRT 5 1 day

Xu et al. (2018)

Synthetic

0.25 L; HRT 5 2 days

Highest power density produced in nonseparator system (66.22 mW/m2), which is 3.9 times higher than the value in the GW system (17.14 mW/m2); biofilm can be severed as the “microbial separator” At C:N 5 2:1, nitrate removal 5 69.3% on 0.583 mA applied current (3.1% higher than normal CW)

5.

Interelectrode distance optimization in vertical upflow CW-MFC Anode: GAC Cathode: GAC

Synthetic wastewater containing phosphate buffer

12.4 L; HRT 5 2 days

6.

Tiered CW-MFC Anode: activated carbon covered stainless steel mesh Cathode: graphite gravel-covered stainless steel mesh Surface flow CW-MFC system using submerged plants and enclosed anodes Electrodes: carbon fiber brush Vertical upflow CW-MFC with and without GW separation Electrodes: carbon felt (Φ 5 10 cm)

Synthetic

5L

Synthetic

CW-MEC for enhanced denitrification in low carbon conditions. Electrodes: graphite granules and rods

7.

8.

9.

Xu et al. (2017)

Srivastava et al. (2018a)

CW, Constructed wetland; GAC, granular activated charcoal; GW, glass wool; IA, intermittent aeration; MFC, microbial fuel cell; PAC, powdered activated charcoal; ROL, radial oxygen loss; HRT, hydraulic retention time; BOD, biological oxygen demand; TEPS, total extracellular polymeric substances; MPD, maximum power density; OLR, organic loading rate; OCV, open circuit voltage; TN, total nitrogen.

Table 13.2 Studies demonstrated enhanced wastewater treatment performance and power output efficiency using CW-MFC. Sr. no.

Wastewater type

Reactor type (operating mode)

Volume

Organic loading rate (kg/m3/D)

COD removal

Maximum power density

References

1.

Synthetic with methylene blue dye Swine

Vertical flow (batch)

5.0 L

0.375

75%

15.73 mW/m2 (anode area)

3.7 L

4.77 6 2.55 (continuous) 0.48 6 0.26 (batch) 0.175

76.5%

9.4 mW/m2 (anode area) (in continuous mode)

Yadav et al. (2012) Zhao et al. (2013)

90%
95%

43 mW/m2 (anode area) CE 5 0.27%
0.45%

Villasenor et al. (2013)

0.302 W/m3 (working volume of the anode)

Fang et al. (2013)

0.852 W/m3 (CE of 1.89%)

Fang et al. (2015) Corbella et al. (2014) Liu et al. (2014)

2.

3.

Synthetic wastewater

Vertical upflow (batch mode, continuous) Horizontal subsurface flow (continuous)

4.

Synthetic with azo dye

Vertical upward (continuous)

5.

Synthetic with azo dye

Vertical upward (continuous)

6.

Urban

Horizontal subsurface flow

7.

Synthetic

8.

Urban

Reactor dimension: 1.15 3 0.47 3 0.5 (L 3 W 3 H, m) Porosity 5 0.4 12.4 L




Reactor scale: 0.3 3 0.5 (Φ 3 H, m) 52.5 L

0.045

Decolorization rate 5 91.24% COD removal 5 85.65% 95.6% decolorization

9.5 6 0.07 g/m2D

69.47%
81.5%




Vertical upward (continuous)

1.4 L

0.016
0.33

81%
95%

Horizontal subsurface flow (continuous)

Cell scale: 0.05 3 0.4 (Φ 3 H, m)

0.124

61%

1.76 mW/m2 (cathode area) with SSM biocathode; 55.05 mW/m2 (cathode area) with GAC-SSM biocathode 36 mW/m2 (anode area)

Corbella et al. (2015)

9.

Contaminated groundwater

Horizontal subsurface flow

10.

Swine

Simultaneous upflow
downflow vertical CW-MFC Vertical upflow CWMFC

Swine

Reactor scale: 2.01 3 0.05 3 0.6 (L 3 W 3 H, m); 8.1 6 0.12 L







1.74 mW/m2 (anode area)

Wei et al. (2015)

0.583 6 0.092

64% 6 4.6%

0.276 W/m3 (anode volume)

8.1 6 0.12 L

0.411
0.854 mg/L

81%

0.168 W/m3 (anode volume) 87.79 mW/m2 (anode area) with the addition of 10% PAC in the anode chamber 0.200 W/m3 (volume of CW-MFC) Internal resistance 5 176.8 Ω


Doherty et al. (2015b) Doherty et al. (2015c) Xu et al. (2016b)

11.

Synthetic

Vertical upward (continuous)

1.5 L

0.484 6 0.063 mg/L

81.4%

12.

Synthetic dye (X-3B) with glucose cosubstrate Synthetic dye (methyl orange) with glucose cosubstrate

Vertical upward (continuous)

12.4 L

0.066

Vertical upward (continuous)




0.0166

Sequencing (batch mode)

2.7 L (smaller substrate size)

0.101

Decolourization rate 5 92.79% (8.86% higher than open circuit) Decolourization rate 5 87.60% (12% higher than open circuit) DMPD degradation rate in anode 5 96.33% (12.37% higher than open circuit) 65.3% 6 3.8% (larger substrate size) 86.7% 6 3.8% (smaller substrate size)

13.

14.

Synthetic

2.9 L (larger substrate size)

8.91 mW/m2 (CE 5 0.22 6 0.18%) —(smaller substrate size); 8.39 mW/m2 (anode area) (CE 5 0.22 6 0.03%) —smaller substrate size

Li et al. (2016)

Fang et al. (2016)

Wang et al. (2017a)

(Continued)

Table 13.2 (Continued) Sr. no.

Wastewater type

Reactor type (operating mode)

Volume

Organic loading rate (kg/m3/D)

COD removal

Maximum power density

References

15.

Synthetic

6.48 L




82.32% 6 12.85%

3714.08 mW/m2

16.

Synthetic

Vertical upward (continuous) Vertical upward (continuous)

2.65 L

0.082 6 0.079

92%

17

Synthetic

2.2 L




TN 90.0%
94.4% NH41-N—64.8% and COD—99%

18

Synthetic

Hybrid horizontal CW-MFC followed by vertical CWMFC Hybrid horizontal CW-MFC

26.12 mW/m2 (anode area) CE 5 1.64% 224 mW/m3

Xu et al. (2018) Yakar et al. (2018) Srivastava et al. (2020)

65.0 L

0.15
0.52

98.3% and 98.9%

CW-MFC, Constructed wetland-microbial fuel cell; GAC, granular activated charcoal; PAC, powdered activated charcoal; CE, coulombic efficiency; TN, total nitrogen.

11.67 mW/m3

Srivastava et al. (2020)

Integration of microbial fuel cell into constructed wetlands: effects, applications, and future outlook 287

of CW-MFC in decreasing the size of the treatment plant where anode occupied only 13.6% volume of the reactor but removed 33% of total COD. Further, Doherty et al. (2015b) examined the effect of electrode distance in electricity generation and found that by decreasing the interelectrode gap, internal resistance decreased by 41% (508 to 300 Ω). Considering the importance of membrane-less CW-MFCs, the upflow
downflow feeding configuration was investigated, which enhanced power density by 70% (Doherty et al., 2015b). Li et al. (2016) and Fang et al. (2016) while treating synthetic textile wastewater containing X-3B and methyl orange, respectively, with glucose as cosubstrate in vertical upflow reactor found that the decolorization rate was 8.86% and 12.37% higher in closed circuit, respectively, as compared to open circuit. Yakar et al. (2018) reported a COD removal of 92% while treating synthetic wastewater (initial COD 5 0.082 6 0.079 mg/L) in a vertical upflow reactor with an achieved maximum power density of 26.12 mW/m2 (anode area) and coulombic efficiency of 1.64%. Liu et al. (2017) has also observed that the performance intensification of CW with integration technology (CW-MFC) also results in lower greenhouse gas production and land footprint, along with energy harvesting.

13.3

Challenges and future perspectives

The treatment efficiency of CW-MFCs has increased significantly after the merger of MFC into CW. The placement of electrode at anode and cathode has a significant impact at an individual level such as electrode at anode act as an artificial electron acceptor and proven for efficient organic removal. On the other hand, electrodes at the cathode act as an electron donor to address pollutant removal such as nitrate (Srivastava et al., 2018a). The potential of anode and cathode still needs to be explored deeply in CW-MFCs for its potential impact. Besides, the main challenge of CW-MFC is to bring it to the field scale, until now, all the studies related to CW-MFCs have been done on the laboratory scale. The significant scale-up challenges, due to which CW-MFC is still on the laboratory scale, are the electricity generation, operational parameters, design configuration, and electrode material and size. In terms of electricity generation, CW-MFC is still behind two order of magnitude from MFCs, which can produce 30 W/m3 of power density from reactor volume of .2 L (Doherty et al., 2015a). Similarly, in terms of electrode positioning, the maintenance of thermodynamically favorable conditions is an essential factor in achieving higher energy production and for the upkeep of kinetics. The internal resistance of the reactor, ohmic, and activation losses is the factor due to which the kinetics (reaction rate) of the CW-MFCs is suppressed (Doherty et al., 2015a; Ramı´rez-Vargas et al., 2018). In addition, the potential of anode and cathode is also an important factor for the maximum extraction of electrons from the organics, which is still not in the focus of CW-MFCs. These all factors mentioned previously hinder the electricity generation in CW-MFCs. On the other hand, if we consider the energy demand for the

288

Integrated Microbial Fuel Cells for Wastewater Treatment

maintenance and operation of CW-MFC, the total energy content of medium strength wastewater is three times higher than the energy requirement (Srivastava et al., 2018b). Therefore the synergy of CW-MFCs has a high impact on the advancement of traditional CW. Furthermore, the anode as an artificial electron acceptor also opens the possibility of a reduction of methane production, which controls greenhouse gas production. The electrogens are a good competitor of methanogenic bacteria, in comparison to methanogens, electrogens dominates the anode portion (Yadav et al., 2018) in bioelectrochemical systems. Moreover, the presence of an artificial electron acceptor at anode also gives the possibility for the manipulation of the microbiology and the processes involved (Yadav et al., 2018). The presence of electrode at anode also gives the possibility for the treatment of industrial wastewater and the recalcitrant pollutants (Srivastava et al., 2018b; Yadav et al., 2018). Despite this, as recommended by Ramı´rez-Vargas et al. (2018), the CW-MFCs need to be further investigated for the potential of anode and cathode for allowing the electrogens for their enhanced kinetics. The monitoring of internal resistance of the CW-MFCs needs to be explored further along with some engineered solutions to reduce the internal resistance. Moreover, the merger of MFC into CW was with the focus to enhance the treatment efficiency of the CW, so despite thinking to enhance electricity generation, more focus should be on optimization and scaling-up.

Acknowledgment SG and AKY acknowledge the financial assistance through the research grant of NASF, (ICAR, New Delhi) (NASF/CA-6031/2017-18). SG heartily thanks AcSIR for PhD support. PS sincerely thanks the financial support received from the University of Tasmania for PhD Study.

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370.

Microbial fuel cell coupled with anaerobic treatment processes for wastewater treatment

14

Suman Bajracharya Water Desalination and Reuse Center, King Abdullah University of Science and Technology, Thuwal, Saudi Arabia

Chapter Outline 14.1 Introduction 295 14.2 Integration of microbial fuel cell in anaerobic digestion 297 14.3 Microbial fuel cell coupling to treat undigested organics in the effluents of anaerobic digestion 299 14.4 Microbial fuel cells coupled anaerobic digestion for nutrient recovery and toxicity removal 303 14.5 Microbial fuel cell coupling in anaerobic digestion as a biosensor for process inhibitors 304 14.6 Outlook 305 References 306

14.1

Introduction

Wastewater treatment facilities at its present stage consume a huge amount of energy; almost 3% of the electricity generated is invested in wastewater treatment plants (WWTPs) in developed countries (Curtis, 2010). At the same time the emission of greenhouse gases from WWTPs has also been increasing with the everincreasing wastewater generations by the growing population (Parravicini et al., 2016). Highly efficient and cost-effective technologies that can offset energy consumption and increasing costs are needed for wastewater treatment to sustain both water and energy security. Bioelectrochemical technologies—comprising microbial fuel cells (MFCs), microbial electrolysis cell (MEC), and microbial electrosynthesis —are emerging wastewater-based sustainable technologies that offer waste treatments with energy savings or even generating energy. Furthermore, bioelectrochemical systems offer new opportunities for multiple ways of waste utilization, for example, the recovery of nutrients, resources, and value-added products from waste (Bajracharya et al., 2017; Rabaey et al., 2010). MFC technologies incorporate the electrochemical interaction of microorganisms with which the electrons generated during the anaerobic degradation of organic matters are transferred to a solid electrode (anode) to produce electricity while Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00014-X © 2020 Elsevier Inc. All rights reserved.

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simultaneously treating the wastes. The electrochemical interaction of microbes uniquely occurs as an extracellular electron transfer to the solid electrode, and such microbes are termed exoelectrogens (Logan, 2008). Electric current is produced when the electrons generated from the biological oxidation at the anode are allowed to flow through an external circuit and recombine with the proton and oxygen at the cathode to form water (Logan et al., 2006). A wide variety of substrates can be used in MFC, including nonfermentable substrates such as acetate (Bond and Lovley, 2003), fermentable substrates such as glucose (Feng et al., 2010), and also inorganic compounds such as sulfides (Rabaey et al., 2006). MFCs produce electric power from the microbial oxidation of a wide variety of substrates that originated from diverse wastewater sources (Pant et al., 2010). Power production in MFCs has been shown by feeding low chemical oxygen demand (COD) wastewater of 21.8 mg/L (Jang et al., 2006) as well as high COD wastewater 127,500 mg/L (Zhang et al., 2009). MFCs have also been studied for electricity production by treating the real wastewater treatment system (ElMekawy et al., 2015; Pant et al., 2010). The power produced from MFC fed with real wastewater ranges up to only several tens of mW/m2 (milliwatt per square meter of electrode surface) (Rodrigo et al., 2007), in contrast, about a few thousand mW/m2 achievable with synthetic effluents (Song et al., 2018). Recovery of energy as electricity from the waste streams along with the removal of waste is the main benefit of MFC technology, which can transform the energyintensive treatment process into energy-neutral or energy positive. MFC technology has prospected as a concept of increasing the efficiency of wastewater treatment and at the same time offers a net energy production (McCarty et al., 2011). The power production has not been upgraded significantly in MFC even though several studies applied advanced nanomaterials in the electrode and improved the reactor configurations (Bajracharya et al., 2016). Low conductivity and low buffer capacity of wastewater effluents treated in MFC are repeatedly referred to as one of the main issues responsible for reduced performances (Rozendal et al., 2008). The developments in materials and reactor configurations of MFCs have brought forward the studies in various sizes of MFCs, treating a few milliliters of wastewater to several 100 L of liquid volumes accommodating MFCs at pilot-scale (Heinemann et al., 2016; Waller and Trabold, 2013). But very limited studies are available on the engineering of MFCs in real/practical application. Studies on largescale MFCs demonstrate the issues of low efficiencies, low power generation, and very high costs compared to the lab-scale MFCs, which turn the MFCs inapplicable in the application (Dekker et al., 2009). Although the power in MFC has not been so attractive, MFCs can still be used in enhancing the wastewaters treatment capacities and sustainability of the treatment system (Oh et al., 2010). In the context of broadening the application of MFC, a number of hybrid applications of MFCs with other technologies (e.g., activated sludge process and anaerobic treatments) have been put forward, and the feasibility to be integrated with the conventional wastewater treatment processes has shown that the integration not only enhanced the waste treatment but also created a selfsustained system (Yoshizawa et al., 2014; Ren et al., 2014). MFCs can be integrated

Microbial fuel cell coupled with anaerobic treatment processes for wastewater treatment

297

with the existing wastewater treatment system to harvest clean electricity (Xu et al., 2016). MFCs can be utilized as a processing unit after the primary treatment or after the anaerobic digestion (AD) process or even as a stand-alone process to remove the organic compounds (Premier et al., 2013). The integrated systems may tackle harsh problems and often become more effective than stand-alone systems (Xie et al., 2014). The advances made on the MFC integrated wastewater treatments, especially the coupling of MFCs to the anaerobic treatment system, appear to be promising in terms of improvement in treatment efficiencies and resource recovery. Here, in this chapter, we reviewed the multiple prospects and progress made in the integration of MFCs in anaerobic treatment systems, especially in AD and highlighted the advantages of the integrated system. This chapter specifically presents an overview of the systems that coupled MFC with AD for practical use. In addition, the challenges to overcome in the system to ensure the adoption of technology by industries are discussed.

14.2

Integration of microbial fuel cell in anaerobic digestion

In conventional wastewater treatment systems the discharge stream still has higher organic content than the set level. A posttreatment step for the effluent quality refining is essential to keep the discharge within the effluent standards (Aiyuk et al., 2004). The anaerobic treatment technologies such as upflow anaerobic sludge blanket (UASB) reactors/anaerobic filters have been in use in anaerobic waste treatment for transforming the organic streams into useful yields. Anaerobic treatment of waste streams can attain net energy production and, at the same time, lower the waste level under the effluent standards. AD is a generally used technology capable of treating the organic wastes at commercial scale and extracting energy from wastewaters as biogas (Schievano et al., 2016). However, AD can recover only 40% of the energy from the organics (Rittmann, 2008). The effluent from AD contains a high amount of biodegradable organics, which required further treatment steps before final discharge. The digestate from AD can also be utilized for producing the fertilizers and conditioning the agricultural soils, but still, many processing steps with energy inputs are required to optimize it for use. In these aspects, MFC, as a posttreatment unit, can treat the undigested waste by further oxidizing the organics from the digestate with the concomitant recovery of energy and nutrient. Coupling of MFC in AD can be suggested to maximize the energy recovery and removal of organics from the effluent (Gao et al., 2014). In particular, air-cathode MFC was demonstrated to have potentially high COD-removal efficiency with a variety of liquid streams (Fradler et al., 2014) and, in addition, to act simultaneously as nitrification denitrification systems (Virdis et al., 2010). An MFC integrated into AD can also help in removing ammonium (nutrient) and recovering it as ammonia at the cathode (Kuntke et al., 2012). At the ambient temperature the bioelectrochemical AD can save the heat input, which is optional for organic waste such as sewage

298

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sludge in cold and moderate regions (Feng et al., 2018). Another important utility of MFC in AD is the use as a biosensor for the online monitoring of operational states of the anaerobic process. Depending on the changes in the voltage signals from MFC, the operational efficiency of the system can be optimized. MFCs can be installed in AD processes in two possible routes—(1) as a standalone application incorporating both anaerobic degradation and bioenergy production resulting a net energy-neutral or energy-positive treatment facility and (2) as an additional polishing unit positioned in series with an anaerobic digester to further reduce the organics in effluent of the anaerobic digester while a stand-alone MFC may not perform at high rate as the AD for the primary treatment of high strength wastewaters. At the lab-scale with synthetic effluent, MFCs are reported to achieve a maximum COD removal rate of 4 kg COD/m3 day (Peixoto et al., 2013). The realistic route of MFC integration in wastewater treatment is to operate on the anaerobic digesters effluent as a final refining step in which the COD and total solids have already been lowered. The MFCs can effectively treat the residual organics in the effluents after AD as substrates (Aelterman et al., 2006). COD level has been lowered in the effluent to ,1 g COD/L after AD, which is a suitable organic load for the operation of MFC (Fradler et al., 2014; Higgins et al., 2013). The concept of integration of MFC in the AD process in the wastewater treatment system is depicted in Fig. 14.1. Electricity production from the oxidation of organic wastes in MFC may not have a substantial economic gain, but the waste removal from the effluent or possible conversion to value-added products in bioelectrochemical processes could be the more attractive aspect of MFC technology integration in waste treatment (Bajracharya et al., 2016). It can also improve the sustainability of the treatment system (Li et al., 2014). However, a full-scale application of MFC has not been realized with real wastewater treatment except one or two test studies at pilot scales. The advances and utilities of MFC integration in AD, as reported in recent literature, are further discussed in the following separate sections.

Low-strength wastewater

Electricity ammonia

Primary settling tank

Wastewater

CH4 H2

High Phosphate wastewater

MFC

Clean water

Membrane filtration Anaerobic digestion

UASB

Struvite precipitation

Fertilizer High-strength wastewater

Figure 14.1 Process flow for a conceptual MFC-centered hybrid process for wastewater refinery. The arrows indicate the water/sludge flow direction. MFC, Microbial fuel cell.

Microbial fuel cell coupled with anaerobic treatment processes for wastewater treatment

14.3

299

Microbial fuel cell coupling to treat undigested organics in the effluents of anaerobic digestion

The effluents from AD usually hold a large amount of biodegradable organic residues [mostly .0.5 g/L volatile fatty acids (VFAs)] (Van Lier et al., 2001). In the AD of domestic wastes an accumulation of VFAs to high level especially in food waste fed AD (Atasoy et al., 2018; Shi et al., 2017) is a major issue which inhibits the methanogenesis and increases the organic load in AD effluent. These unrecovered organics could be converted to electrical energy using MFC by using MFC as a downstream refining step in the AD process. The coproduction of electricity, while cleaning up the AD effluent, can potentially compensate a fraction of the energy consumption in the digestion system. Several literatures have demonstrated the feasibility of the treatment of the residual constituents in AD effluent by integrating MFC to produce permitted final quality effluent. Table 14.1 presents a summary of several options adopted in the treatment of effluents from AD and leachates from solid waste for the production of electricity. Zhang et al. (2009) reported the treatment of high strength molasses wastewater in a system consisting of a UASB reactor, an MFC, and then a biological aerated filter (BAF) sequentially (UASB MFC BAF), which also generates electricity continuously. In this study, MFC was used for generating current mainly from the oxidation of sulfide generated in the UASB reactor due to sulfate reduction. When the UASB-MFC-BAF system was treating the 20 times diluted raw molasses wastewater, the COD removing efficiencies of the system remained around 30% and when the feed was switched to five times diluted wastewater, the efficiency of the system was highly improved, reaching almost 60.0% (Zhang et al., 2009). Feeding of raw molasses wastewater to the UASB reactor dropped the COD removal efficiency radically because of the suppression induced by substrate inhibition. Higher COD removal efficiency achieved in the UASB MFC BAF system than the conventional process for molasses wastewater treatment and energy recovery was also achieved. In a similar setting an integration of MFC with biohydrogen producing biofermentor was reported in Sharma and Li (2010) for the enhancement of bioenergy harvest from wastewater treatment. The combination of hydrogen-producing biofermentor and MFC enhanced the COD removal, reaching 59% 71% from the artificial wastewater containing 2 10 g/L of glucose, and the total energy recovery of 29% was achieved in the combined hydrogen and electricity production from wastewater. Ge et al. (2013) reported the use of tubular MFC for treating the digested sewage sludge containing 16.7 11.4 g/L of total COD from anaerobic digester of municipal wastewater treatment with an average of 36.2% additional COD removal. The MFC produced electricity on the treatment of digested sewage sludge at the COD removal rate of 1.86 kg/m3/day and attained a maximum power output of 3.2 W/m3, which corresponds to 2.6% of coulombic efficiency (Ge et al., 2013). As the food waste leachate produced after the hydrolysis and acidogenesis stages of AD is rich in VFAs, MFC was used to produce electricity directly from the VFAs obtained in

Table 14.1 Treatment efficiency of microbial fuel cell (MFC) coupled anaerobic digester reported in several literatures. References

Feedstock

System

COD removal

Current density (A/m3)

Power output (W/ m3)

Zhang et al. (2009)

Molasses wastewater treatment, initial COD 127.5 g/L

UASB MFC BAF

4.95 A/m2 (MFC sulfide oxidation)

1.41 W/m2

Sharma and Li (2010)

Artificial WW with glucose 2 g/L COD 10 g/L COD Artificial wastewater 2.4 g/L COD with 0.4 g/L sulfate

H2 biofermenter 1 MFC

COD removal 53.2% Sulfate removal 52.7% Color removal 41.1% 71% 59% 82% COD removal 70% sulfate removal 36.2% 6 24.4% (in MFC only)

0.4 0.6 mA

888.9 6 10.5 mW/m2

34

3.2

G

G

G

G

G

3.493 4.2

G

Zhang et al. (2012)

Ge et al. (2013)

Li et al. (2013) Tugtas et al. (2013) Inglesby and Fisher (2012)

Municipal WW 16.7 6 11.4 g total COD/ L Food waste leachate 1 g COD/L Landfill leachate 4.64 7.02 g COD/L Arthrospira maxima biomass with 0.5 3 g DW/L/day OLR

AD 1 MFC

AD 1 tubular MFC AD 1 2 chamber MFC

87% 92%

2.2

0.455

USAB 1 2 chamber MFC

.95%

8.3

2.5

AD 1 recirculation loop 2 chamber MFC with packed graphite granules

69% 6 6% (at 0.5 g/L/day OLR)

16.8 6 0.9 18 6 0.9

7.8 6 0.9 10.2 6 1.3

Inglesby and Fisher (2013) Fradler et al. (2014) Schievano et al. (2016)

Liu et al. (2019)

A. maxima biomass with 1.31 5.38 g total COD/ L Wheat feed pellets 0.572 g soluble COD/L/day The mixture of swine manure and rice bran 124 6 8.1 g COD/kg fresh matter

AD 1 2 chamber MFC with packed graphite granules BioH2 1 AD 1 4 MFCs in series Dark fermentation 1 AD 1 single chamber aircathode MFC

The mixture of VFAs at 15.3 g organic C/L resembling food waste

Anaerobic fermentation liquid catalytic fuel cells (AF-FC)

52% 85%

13

5.8

35.1% 44%

2.2

5.5 W/m3

22% COD removal 51% VFA removal 19% Total nitrogen removal 93.1% TOC removal in 60 h . 82% TN removal

0.738 Wh/kg raw feed

1.2 mW/cm2

AD, Anaerobic digestion; BAF, biological aerated filter; COD, chemical oxygen demand; MFC, microbial fuel cell; WW, wastewater; OLR, organic loading rate; UASB, upflow anaerobic sludge blanket; VFAs, volatile fatty acids.

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Integrated Microbial Fuel Cells for Wastewater Treatment

partial AD of municipal solid waste, majorly constituted by food waste (Li et al., 2013). Li et al. (2013) reported that the combination of partial AD and MFC for food waste leachate treatment resulted in 87% 92% of COD removal with 0.45 W/ m3 of power production. In a similar study for landfill leachate treatment using the combination of AD and MFC, more than 95% of COD was removed and also produced electricity with 2.5 W/m3 of power density (Tugtas et al., 2013). An MFC is having packed graphite granules as electrodes were introduced in a recirculation loop of AD effluent to treat and produce electricity from the residual organics generated after AD of cyanobacteria biomass (Inglesby and Fisher, 2012, 2013). The integrated system of AD and MFC showed increase of the COD removal reaching up to 69% 6 6% of COD removal at a hydraulic retention time (HRT) of 4 days and organic loading rate (OLR) of 0.5 g DW/L day which produced 16.8 6 0.9 A/m3 of current density and 7.8 6 0.9 W/m3 power density (Inglesby and Fisher, 2012). Overall, methane yield was increased by 50% with the integration of MFC in AD, which reached 350 mL CH4/g VS at 0.5 g DW/L/day OLR. A similar second study by Inglesby and Fisher (2013) reported the 52% 85% of COD removal efficiency at the COD removal rate of 1.7 13.4 kg COD/L/ day and current output of 13 A/m3. Fradler et al. (2014) used four 0.25 L MFCs in series to refine the effluent generated after biohydrogen and biomethane production processes operated in sequence fed with wheat feed pellets. AD sludge inoculated MFC operated at an HRT of 33.3 hour with six different organic loadings [0.036 6.149 g soluble chemical oxygen demand (sCOD)/L day] removed 35.1% 4.4% of COD and the COD removal efficiency diminished at higher OLRs. When the OLR was 0.57 g sCOD/L, the MFC produced a power density of 3.1 W/m3/day resulting in the highest coulombic efficiency of 60% (Fradler et al., 2014). In similar study the treatment of a mixture of swine manure and rice bran in an integrated system consisting of a dark fermentation for hydrogen production, an AD for methane production, and then MFC for electricity production resulted 22% of total COD removal, 51% VFA removal, and 19% total nitrogen removal with the production of 0.738 Wh electrical energy/kg of raw feedstock (Schievano et al., 2016). The coupling of MFC and anaerobic membrane bioreactor having fluidized-bed can remove high levels of COD (92.5%) and total suspended solids ( . 99.0%) (Ren et al., 2014). A new approach to utilize fermentation products as electrical energy is shown by combining anaerobic fermentation and liquid catalytic fuel cells, which attained 93% of total organic carbon (TOC) removal from the mixture of VFAs resembling wastewater from food waste (Liu et al., 2019). Compared to the conventional anaerobic processes, electrochemical intervention in the anaerobic fermentation exhibits 34% coulombic efficiency for food waste utilization (Liu et al., 2019). Multiple examples of coupling of MFC with AD from literature spotlight the potential of MFC as a downstream refining unit to meet the discharge quality standard of disposable wastewater at the same time recover electrical energy from AD effluent. In the case of an increase in VFAs in the effluent, inhibition of the AD process may frequently occur. At such event, an integration of MFC in AD can avoid the VFA accumulation in the digestion liquor. MFCs can mainly be

Microbial fuel cell coupled with anaerobic treatment processes for wastewater treatment

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advantageous as a buffering system to maintain and recover the AD process whenever AD suffers from inhibition (Cerrillo et al., 2017). The low electrical power production and low economic value of the electricity limit the large-scale application of MFC for power production (Li and Yu, 2014; Logan and Rabaey, 2012). Furthermore, the scaling-up of MFC is not attainable by just expanding the size of the electrodes and reactor volume or in another case, by linking multiple MFCs together in series or in parallel to increase power harvesting; the reason is due to the non-linear behavior of MFC systems (Wang et al., 2015). The coupling of MFCs in existing anaerobic processes could be the most practical aspect.

14.4

Microbial fuel cells coupled anaerobic digestion for nutrient recovery and toxicity removal

Ammonia, sulfides, and VFAs can accumulate in AD broth, which can inhibit the methanogenesis process and also result in reactor instability. Treatment of AD effluent in MFC could lower the inhibitory concentration of ammonium. In an MFC having a cation exchange membrane the accumulated ammonium in anolyte would transfer to the cathode side, which results in ammonia removal in continuous and selective fashion from the digestate. This approach has been applied in a number of studies using several waste streams such as synthetic wastewater, AD digestate, urine, and effluent from sewage sludge treatment (Desloover et al., 2012; Kim et al., 2015). In this process the potential difference developed between anode and cathode of MFC creates a net flux of ammonium ions from the waste stream at anode compartment to the counter stream at the cathode compartment (Cheng et al., 2013). Principally, when the ammonium ions reach the cathode compartment crossing the cation exchange membrane, an alkaline catholyte of pH . 9.2 transforms ammonium to volatile ammonia due to the acid dissociation constant (pKa value) of 9.25 (at 25 C) for ammonium and removed from the catholyte by the volatilization. The recovery of ammonia is carried out by dissolving in acids separately. The use of MFC in ammonia stripping from anaerobic digester is an attractive approach for ammonia inhibition and enhancing AD performance; in addition, it recovers nutrients from the anaerobic digestate and generates electricity. Desloover et al. (2012) observed the ammonium transfer efficiency of 96% with 2 NH1 4 flux of 120 g N/m /day across the cation exchange membrane in MFC and produced electrical energy of 5 kWh/kg N of removed. The ammonium concentration in the digestate was significantly lowered from 2.1 to 0.8 1.2 g N/L (Desloover et al., 2012). The bioanode facilitates the further oxidation of VFAs in the digestate and thereby avoids VFA accumulation and its inhibitory effect in the digestion process. Inglesby and Fisher (2012) also demonstrated the removal of ammonium in a lab-scale MFC coupled advanced flow-through anaerobic digester. A nitrogen-rich Arthrospira maxima cyanobacterial biomass having 77 125 g N/kg total solid was fed to the AD. Ammonium level was lowered from 654 to 436 mg/L by integrating MFC in a recirculating side stream of AD, which accounts for up to

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33% of ammonium removal (Inglesby and Fisher, 2012). The authors reported that the methane yields in AD process were considerably improved by 27% after MFC integration yield increased from 136 to 173 mL CH4 from one gram of volatile solid and the energy recovery in the MFC coupled AD system increased by 36% (from 22% to 30%) (Inglesby and Fisher, 2012). Almost 78% of total ammonia nitrogen (TAN) removal by the subsequent treatment in MFC was also exhibited by Kim et al. (2015) from the effluent of AD digesting swine wastewater containing slightly higher than 4 g TAN/L. During the AD process, only 5.8% of ammonia is removed, and the accumulation of TAN lower the COD/TAN ratio, which resulted in decrease in COD removal in AD. With the coupling of MFC the TAN level was reduced, and further COD removal was achieved accounting for up to 90% of COD removal from the combined system (Kim et al., 2015). These studies support the beneficial role of MFC integration in AD of high nitrogen containing feedstock such as food waste, poultry, and swine wastewater. Besides the removal of ammonia, AD digestate recirculation through a cathode compartment of MEC enables the removal and recovery of phosphorus. In a recent study, a concept of phosphorus removal from the digestate of an anaerobic digester treating the secondary sewage sludge was achieved by processing in a MEC comprising a cation exchange membrane and a large surface fluidized-bed cathode (Cusick et al., 2014). Cusick et al. (2014) accustomed the applied voltage and the loading rate of digestate to keep the digestate fluidized and maintain alkaline (pH 8.7) by the cathodic proton-consuming reaction that enabled the in situ formation/ precipitation of struvite (magnesium ammonium phosphate hexahydrate) at the cathode from the AD digestate. The in situ struvite formation with an applied voltage of 1 V recovered up to 82% of soluble phosphorus from the digestate (Cusick et al., 2014). This study illustrates the synergetic usage of MFC-based technologies with AD for progressing in nutrient recovery. One of the major problems to be tackled in AD is the toxicity of components present in various industrial wastewaters. A common problem here is related to the presence of sulfate, which is reduced by sulfate-reducing bacteria (SRB) to sulfides in the anaerobic process. When sulfate is present, the SRB become active and compete with methanogens for the electron donor. Besides, more importantly, sulfides are toxic and can also inhibit the process. While feeding the sulfide containing effluent from AD to an MFC, the sulfide was biologically and chemically oxidized to elemental sulfur, with a small amount of sulfate and at the same time generate electricity (0.4 0.6 mA current) (Zhang et al., 2012). COD removal efficiency increased from 68% to 82% after integrating MFC to AD while treating sulfate containing 2.4 g/L COD wastewater.

14.5

Microbial fuel cell coupling in anaerobic digestion as a biosensor for process inhibitors

High VFAs and ammonia levels can trigger the process instability in AD (Chen et al., 2008; Appels et al., 2008). High levels of VFAs accumulation, mainly

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propionate, can impede microbial activities that might lead to process failure in AD (Nielsen et al., 2007; Yuan and Zhu, 2016). MFCs integration in AD has shown to reduce the inhibitory VFAs level and facilitates the operation of AD to enhance the removal of organics in the effluent with the concurrent recovery of nutrient as ammonia (Cerrillo et al., 2017). Cerrillo et al. (2017) reported that the MFC attained 50% COD removal during AD inhibition and 31% ammonia removal at the rate of 11.19 g N/m2/day. The propionic-acetic acid ratio is a key process indicator in AD (Marchaim and Krause, 1993). An in situ real-time sensor for VFA level monitoring has been a convenient approach to monitor the AD process. MFC could be used for online monitoring of the VFAs, including acetate and propionate concentrations; the electrical indicators such as electrode potentials, current cell, and voltage can be used for detecting the VFA concentrations (Kaur et al., 2013, 2014). Liu et al. (2011) showed a prototype of MFC biosensor as a useful tool to regulate the processes in AD by deploying an MFC in the effluent recirculation loop of a UAFB (upflow anaerobic fixed-bed) reactor. The voltage developed between anode and cathode of the MFC was dependent on the OLR of the UAFB reactor in such a way that a rise in organic loading from 1.0 to 2.5 g COD/L/day escalates the voltage in MFC from 20 mV to 80 mV when operated with 800 Ω external resistance and viceversa (Liu et al., 2011). The authors also proposed to use MFC sensor signaling to monitor the other AD process parameters such as pH and biogas production rates to alert for counteractive action in time. Overall, MFC signaling is useful in controlling the AD process, and it provides a cost-effective and relatively less-risky mean of process monitoring, which can also be highly compatible with the fullscale operation of AD.

14.6

Outlook

MFCs have been recognized as promising systems to supplement the AD to enhance the elimination of organic load in the effluent thereby improving the effluent quality to meet the discharge standard. The coupling of MFC in AD can also aid in removing the accumulation of inhibitory levels of VFAs, enhancing the product yields, and recovering nutrients. The MFC coupled AD systems have achieved high organic waste treatment efficiencies and also maintain the process at optimum. The integration of MFC technology in AD potentially turns the AD process more attractive, robust, and sustainable for wastewater treatment. The successful cases from the lab-scale MFCs integrated with AD are remarkable, but no successful real-scale application of MFCs has been achieved so far. A promising AD-MFC integration in real scale will largely be governed by the scalability and functioning of MFC at large scale. Reactor size enlargement and stacking of multiple MFCs are the popular strategies for scaling-up, but these approaches are associated with issues such as ionic short circuit and cell voltage reversal due to non-linearity in electrical properties of MFC technology. Stacking of MFCs by connecting the modules of

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multiple MFCs in series has shown improvement in waste treatment and maximum voltage output. Further research should focus on the multiple possibilities of stacking and should also address the challenges associated with the upscaling and stacking of MFCs. Cost remains another factor for the commercial application of integrating MFC technology. MFC integration to a real scaled-up anaerobic treatment facility may require a high investment on reactor materials to treat a large throughput of the influent wastewaters and gases. Low-cost but effective, the reactor materials have long been in research, and still, the quest is ongoing. Overall, this chapter highlights the merits of the coupling of MFC with existing anaerobic wastewater treatment processes; the increasing research interest in this coupled system has shown multiple benefits in terms of effective treatment, resource recovery, and sustainability aspects. MFCs have been proven to be highly practical when coupled with AD particularly as an effluent polishing unit to improve the effluent quality, either by lowering its organic matter content or by removing nutrients. In the context of bioprocess enhancement, integration of other MFC technology based applications such as microbial electrolysis, microbial electrosynthesis, and electro-fermentation are also promising, and synergistic application should also be studied with the existing anaerobic wastewater treatment processes.

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Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater

15

Thangavel Sangeetha1,2, Chellappan Praveen Rajneesh3 and Wei-Mon Yan1,2 1 Department of Energy and Refrigerating Air-Conditioning Engineering, National Taipei University of Technology, Taipei, Taiwan, ROC, 2Research Center of Energy Conservation for New Generation of Residential, Commercial, and Industrial Sectors, National Taipei University of Technology, Taipei, Taiwan, ROC, 3School of Biomedical Engineering, College of Biomedical Engineering, Taipei Medical University, Taipei, Taiwan, ROC

Chapter Outline 15.1 Introduction 15.1.1 15.1.2 15.1.3 15.1.4 15.1.5 15.1.6

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History of beer 313 Brewing process and wastewater 314 Brewery waste and beer wastewater treatment 315 Bioelectrochemical systems for beer wastewater treatment 317 Anaerobic digestion of beer wastewater treatment 318 Hydrogen production in anaerobic reactors with beer wastewater 318

15.2 Integrated microbial electrolysis
anaerobic digestion for beer wastewater treatment 319 15.2.1 Background 320 15.2.2 An experience of scaling up of the novel microbial electrolysis
anaerobic digestion reactor 322 15.2.3 Overall summary of the experience 341

Acknowledgments References 342

15.1

342

Introduction

15.1.1 History of beer Beer is an ancient beverage produced by humans during at least the fifth millennium BCE in Iran (almost 7000 years back) and also in ancient Egypt and Mesopotamia and spread throughout the globe. Any cereal can undergo fermentation with yeasts to form beer-like drinks and thus were produced after any tribe or culture had started agriculture. These discoveries reveal the earliest records of beer production to date. In Mesopotamia the oldest evidence of beer is a 6000-year-old Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00015-1 © 2020 Elsevier Inc. All rights reserved.

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Sumerian tablet reporting that people consumed a drink using reed straws (Legras et al., 2007). Beer was manufactured from barley. Chinese history illustrates that beer was produced before 5000 years using barley and other grains. Beer was even popular in Neolithic Europe, before 5000 years and was brewed on a domestic scale. Before the Industrial revolution, beer was made and sold on domestic levels, but during and after the revolution the production moved to industrial scales. Later on, the inventions of instruments such as hydrometers and thermometers rapidly changed the brewing process by providing extensive knowledge of the mechanisms and products (Wang et al., 2016). Nowadays, the brewing industry is a highly commercialized global business, with several companies, countless small and medium producers. Almost 1300 million hectoliters of beer were produced globally in 2016, with China being the world’s largest beer producer with more than 440, followed by the United States with 217 million hectoliters (FAOSTAT, 2017; Arantes et al., 2017).

15.1.2 Brewing process and wastewater Beer is the fifth most consumed beverage in the world behind tea, carbonates, milk, and coffee, and it continues to be a popular drink with an average consumption of 23 L/person/year. The brewing industry has an ancient tradition and is still a dynamic sector open to new developments in technology and scientific progress. The manufacturing process of beer from cereals in a brewery is named brewing or beer production, which includes several steps such as malting, mashing, milling, lautering, boiling, fermenting, conditioning, filtering, and packaging. Malting is the step where barley is prepared for brewing and has three steps (Fillaudeau et al., 2006). First, the grains are soaked for almost 40 h. Then they are spread on the floor of the germination room for 5 days. Then finally, the germinated grains are put in a kiln and dried at very high temperatures for several hours. After this the grains are called malt, and they will be mashed for the extraction of the sugars. Mashing is the step where the starches are released during the malting stage and are converted to sugars for fermentation. The milled grain is mixed with hot water in large vessels into a cereal mash, where the enzymes in the malt convert starch to saccharides at 140 C
160 C. A sugar-rich liquid is formed, which is strained out of the vessel; it is called lautering. After this the liquid is boiled in copper kettles, and herbs are added to decide the flavor, color, and aroma of the beer. Then the liquid is clarified in a vessel called a “whirlpool,” where solid particles are separated out. Then the liquid is cooled to 60 C, and yeast is added to initiate the fermentation process where the sugars are converted to alcohol, carbon dioxide, and other components. Then the fermented liquid undergoes conditioning, where the beer ages to give a smooth flavor. After undergoing conditioning for a week to several months, the beer will be filtered and packed for commercial purposes (De Francesco et al., 2018). Breweries are major contributors to the economy of a country, but they are considered as consumers of water and a significant source of environmental pollution. In spite of improvements, water and energy consumption, waste, wastewater generation, by-products, and air-pollutant emissions are still

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 315

Table 15.1 Physicochemical parameters of the real beer wastewater. Parameter

Value

COD pH Ammonia BOD SS

2106
2250 mg/L 6.6
6.8 52
60 mg/L 1285
1540 mg/L 36
450 mg/L

COD, Chemical oxygen demand; BOD, Biological oxygen demand. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Gao, L., Wang, L., Cui, M., et al., 2017. Energy recovery evaluation in an up flow microbial electrolysis coupled anaerobic digestion (ME-AD) reactor: role of electrode positions and hydraulic retention times. Appl. Energy 206, 1214
1224.

remain bottlenecks. Beer wastewater (BW) is a form of agro-industrial wastewater. It is estimated that for the production of 1 L of beer, 3
10 L of waste effluent is generated. Wastewater is produced in almost all the stages of beer production and contains organic matter and suspended solids, which require treatment prior to their discharge into water bodies (Oliveira et al., 2019). The chemical oxygen demand (COD) of BW can be almost 32,000 mg/L. The characteristics of the BW analyzed in our research studies are listed in Table 15.1. The wastewater produced by cooling and washing units has high COD but is nontoxic. Wastewater from a brewery plant may be discharged in several ways such as (1) direct discharge into a waterway (oceans, rivers, streams, or lakes); (2) discharge into a municipal sewer system; (3) discharge into a waterway or municipal sewer system after the wastewater has undergone some pretreatment; and (4) into the brewery’s own wastewater treatment plant. Traditional treatments, such as aerobic sequencing batch reactors, require high energy input and are thus costly. New approaches for wastewater treatment, which not only reduce cost but also produce useful side products, have recently received increasing attention (Chowdhary et al., 2018).

15.1.3 Brewery waste and beer wastewater treatment Waste and wastewater management and disposal are vital cost-consuming factors and significant aspects of breweries. There are several treatment methods to which this wastewater has been subjected to, and the treatment of such wastewater in breweries generally includes processes that are applied to all other industrial wastewaters such as physical, chemical, and biological methods. BW is highly organic in nature as different kinds of organic wastes are being generated in almost all the stages of brewing, and it needs to be treated before discharge into the environment. If the brewery does not discharge the effluent into the municipal sewer, then primary and secondary treatment methods are needed. However, if the brewery has been permitted to discharge into a municipal sewer, then pretreatment methods are required to fulfill the laws and to reduce its impacts on the municipal wastewater

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Integrated Microbial Fuel Cells for Wastewater Treatment

treatment plant (Arantes et al., 2017). Some treatment and management processes that are being adopted in breweries have been described next.

15.1.3.1 Physical treatment BW has been subjected to various physical wastewater treatment processes such as screening, grit removal, and sedimentation. The major solid wastes in BW are the spent grains, kieselguhr sludge, glass bottle pieces, bottle caps, waste labels, floating plastic items, etc. These wastes are usually screened and removed from the wastewater. Flow equalization is adopted to equalize or consolidate the wastewater for holding in tanks before passing on that water for further downstream treatment processes or introducing it directly into the municipal sewage system (Dizge et al., 2018). After the screening, heavy solid waste particles such as grit, sand, and stones may settle down to the tank bottom. They are removed by sedimentation or settling process. Physical forces are applied to eliminate contaminants such as coarse solid particles, suspended particles, and sediments. Sedimentation is performed to remove the suspended particles, and general sieving may also be done to remove the solids (Tsai et al., 2008).

15.1.3.2 Chemical treatment processes Compared to all other chemical treatment methods, adjustment of pH, coagulation, and flocculation are commonly followed in breweries. The pH of wastewater is normally between 2 and 12 and should be maintained in the range of 6
9 to conserve the microbes for further treatment processes. Caustic soda and nitric acids are used for sanitizing purposes, but waste carbon dioxide from fermentation can be utilized as a cheap acidifying agent, thus reducing acid usage. Neutralization can also be done using some acids, but this may not be cost-effective for the treatment plants. This may be followed by disinfection or biological methods. Coagulation and flocculation are done to remove colloidal material and decolorize the wastewater. Nevertheless, physical and chemical treatment methods have drawbacks related to cost, secondary pollutant generation, and the requirement of more energy and power (Simate et al., 2011).

15.1.3.3 Biological treatment methods Biological treatment methods adopted for BW can be generally classified into two major divisions as aerobic and anaerobic treatments. Compared to other treatment methods, biological methods have certain advantages such as (1) treatment technology is traditional and well understood; (2) enhanced efficiency in terms of organic content removal; (3) cost-effective; and (4) environment friendly and safe. A common aerobic method such as activated sludge process is performed by allowing wastewater flow into the aerated tank and agitating the sludge to get a consortium of fungi, algae, bacteria, etc. The vigorous mixing of the biomass enhances oxygen supply, thus facilitating organic matter uptake and microbial reproduction. Trickling filters, rotating biological contactor process, and biotowers have also been

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 317

employed for the treatment of BW. Their overall treatment efficiency was found to be 40%
50% (Simante and Hill, 2015). Anaerobic treatment composes processes that occur in an environment devoid of air or elemental oxygen. It is characterized by the biological conversion of organic compounds by anaerobic microbes into biogas, which comprises 60%
75% methane and 25%
40% carbon dioxide and traces of other gases such as hydrogen, hydrogen sulfide, and nitrogen. The most significant anaerobic treatment processes that are prevalent in breweries are upflow anaerobic sludge blanket (UASB), fluidized bed reactors, and membrane bioreactors (Hidalgo and Marroquı´n, 2019).

15.1.4 Bioelectrochemical systems for beer wastewater treatment The global energy crisis is emphasizing the researchers around the world to invent and discover alternative, renewable energy sources (Wang et al., 2019; Sangeetha et al., 2019). Recently, bioelectrochemical wastewater treatment has emerged as a potentially interesting technology for the production of energy from wastewaters. Microbial fuel cells (MFCs) and microbial electrolysis cells (MECs) are two common types of a bioelectrochemical system (BES) (Wang et al., 2018a,b). They are based on the use of electrochemically active microorganisms, which transfer electrons to the anode electrode while they are oxidizing the organics in the wastewater. MFCs and MECs have been widely used by researchers for the treatment of BW as it was a good choice due to its organic and less harmful nature (Wang et al., 2018c,d). One of the pioneers’ works with MFC was by Feng et al. (2008), where they had employed BW as a substrate in single-chambered air cathode MFCs, with carbon cloth electrodes (both anode and cathode). They had reported 63 mW/m2 electricity generation during such wastewater treatment, which was lesser to the electricity generated during domestic wastewater (i.e., 150 mW/m2). Following them, Wen et al. (2009) had operated single-chambered air cathode MFCs for the treatment of BW. They used carbon fiber as anode and stainless steel net as a cathode. The MFCs displayed a maximum open-circuit voltage of 0.578 V and a power density of 264 mW/m2. Polarization curves were also used for estimating the internal loss, but eventually, the results demonstrated that BW was feasible and stable for producing bioelectricity through MFCs (Wen et al., 2010a,b). MECs are electrochemical transducers that produce hydrogen by the synergizing ability of electrogenic bacteria to oxidize organic matter with the hydrogen evolution reaction at the cathode, utilizing the anode as the electron acceptor. MFCs can be modified into an MEC by adding a small supplement of electricity at the cathode to produce products such as hydrogen gas, thus offering additional advantages (Adekunle et al., 2019). Very recently, BW has been used as a substrate in MECs for the treatment and production of hydrogen gas (Kadier et al., 2014). BW is a suitable substrate option to produce hydrogen gas and electricity as it has high organic content, is generated in large amounts globally, and requires treatment prior

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Integrated Microbial Fuel Cells for Wastewater Treatment

to environmental discharge. Its composition includes sugars from the brewing process and is a viable fermentable substrate for microbes (Arantes et al., 2017). Sosa-Herna´ndez et al. (2016) had employed the spent yeast from the BW as a substrate for MEC. A notable amount of current (220 A/m3) and hydrogen yield (HY) (2.18 LH2/day/L reactor) were reported to be produced.

15.1.5 Anaerobic digestion of beer wastewater treatment Anaerobic digestion (AD) is the conversion of organic matter to methane-rich biogas through a series of interlinked processes such as hydrolysis, fermentation, acetogenesis, and methanogenesis. The end products of AD are usually methane, carbon dioxide, hydrogen, and water. Normally, 70% of the biomass gets converted to methane, whereas the remaining 30% is hydrogen. Many researchers have employed BW for biogas production in AD reactors (Enitan et al., 2015; Xu et al., 2013). Nevertheless, a pioneering research of successfully integrating UASB reactors with absorbers for the treatment of BW was done by Rao et al. (2007). They achieved 96% of COD reduction and methane yield (MY) of 0.32 m3 CH4/g of reduced COD. The study of BW in AD through a field-scale expanded granule sludge blanket reactor to produce biogas through anaerobic reactions was conducted by Caliskan et al. (2014). They reported that 45 m3 of methane was produced/ton of malt waste. Anaerobic sequencing batch reactor was used to treat BW. Beer lees were anaerobically digested, and valuable MY of 5300 mL/day was produced, and the spent lees were used as compost and fertilizer (Sun et al., 2019). Anaerobic treatment of BW for enhanced biogas recovery was performed by Manyuchi et al. (2018), and biogas with a methane composition of almost 70% was achieved. Thus AD has achieved enormous success in BW treatment and biogas production.

15.1.6 Hydrogen production in anaerobic reactors with beer wastewater New techniques for the production of hydrogen by AD processes were attempted by many researchers. Sinbuathong et al. (2015) used dark fermentation to produce hydrogen from BW. They suppressed the methanogens by temperature and pH and induced hydrogen production. But they reported that the gas acquired contained almost 60% of methane and 20% of carbon dioxide rather than hydrogen. BW was used as a substrate for hydrogen production in an anaerobic environment, and biogas with 65% H2 was produced (Golub et al., 2014). The optimum temperature, pH, and BW concentration were evaluated for maximum hydrogen production. All these studies had indicated that practical difficulties were more as the conditions were hard to maintain, and new integrated reactors need to be designed and operated for unhindered and improved hydrogen generation along with methane production and wastewater treatment.

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 319

15.2

Integrated microbial electrolysis
anaerobic digestion for beer wastewater treatment

Considering the abovementioned requirements, there has been a crucial need to integrate significant reactors for the enhanced production of methane, hydrogen, bioelectricity, and for the overall treatment of BW. Coupling of two significant reactors such as MEC and AD can supply abundant hydrogen and carbon dioxide for enhanced methane generation. Electrochemical methanogenesis is more controllable and stable compared to conventional anaerobic methods. This technology is relatively new, and scale-up studies using single-chambered MECs have been performed in our previous studies (Cai et al., 2016a,b). Rapidly developing bioelectrochemical technology is surely a promising platform for methane (CH 4) and hydrogen (H2) production. The addition of an external electrochemical system in an anaerobic reactor can be used to enhance microbial metabolism and wastewater recycling. The microbes in the wastewater utilize the organic matter and oxidize it into H2 and CO2. With a small addition of voltage to the MEC, electromethanogens can use electrons, H2, and CO2 and reduce them into CH4 (Arvin et al., 2019; Cerrillo et al., 2018). Thus research work headed to the integration of ME and AD reactors for in situ conversion of organic matter into CH4 and to treat the BW is promising. This attempt has been considered the novel in two aspects: (1) the construction of the unique combined ME
AD reactor and (2) the employment of BW in this integrated reactor. A schematic representation portraying the importance of the integration is in Fig. 15.1.

Figure 15.1 Significance of MEC-AD integration. AD, Anaerobic digestion; MEC, microbial electrolysis cell.

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Integrated Microbial Fuel Cells for Wastewater Treatment

15.2.1 Background 15.2.1.1 China-Global beer hub Production and consumption of beer in China have occurred form ancient, with recent archeological findings showing that Chinese villagers were brewing beertype alcoholic drinks as far back as 7000 BCE on small and individual scales. Modern beer brewing was not introduced into China until the end of the 19th century when Russians, Germans, and Czechs established breweries in Harbin. China is one of the top five largest beer-producing countries in the world, with year-round production of 492.19 million hectoliters and satisfying more than 25% of the global beer demand. It has more than 200,000 breweries, and for the production of 1 m3 of beer, 10
20 m3 of water is consumed, and 90% of the water is discharged into sewer systems (Li et al., 2019). Harbin city in the northeast of China is a major beer-producing center with many beer industries and the most famous one being the Harbin Beer and the oldest brewery in China (Xie et al., 2016).

15.2.1.2 Significance and application prospects of the reactor Beer industry wastewater is considered to be compatible with methane generation by anaerobic treatment due to its food-derived nature, and it is also rich in organic contents such as carbohydrates, protein, and starch (Caliskan et al., 2014). Traditional treatment methods are expensive, energy-consuming, and less effective compared to anaerobic treatment. Biohydrogen production by AD has been developed as a response to global interest in the use of hydrogen as a clean and efficient energy carrier. It has been employed for the treatment of wastewater by the destruction of pathogens and the generation of bioenergy in the form of biohydrogen and methane and has been recognized as a more controllable and sustainable method. Moreover, it has been proved that methanogens could be controlled or affected through the cathode recovery niche. It gives the great potential to accelerate and adjust methane generation in a conventional AD system (Liu et al., 2016a).

15.2.1.3 Unique advantages of microbial electrolysis
anaerobic digestion reactor over conventional technologies Compared to conventional AD, the coupled reactor could offer some specific advantages. As AD and MEC are integrated, the possibility to remove the residual organics contained in the effluent of a conventional anaerobic digester with such a BES is very effective. Such a new reactor is advantageous than the conventional digesters in the next mentioned ways.

15.2.1.3.1 A complete treatment of a wide range of wastewaters Various types of organic and inorganic wastewaters can be effectively treated in an MEC reactor. So, when a pair of electrodes are inserted into an anaerobic reactor to form an ME-combined AD system, not only high concentration organics in wastewater can be degraded effectively via anaerobic fermentation, but also electrode

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 321

reaction of MEC can be used for advanced treatment of a wide range of wastewaters. The addition of an external electrochemical system in an anaerobic reactor can be used to enhance microbial metabolism and wastewater recycling. The microbes in the wastewater utilize the organic matter and oxidize it into H2 and CO2. With a small addition of voltage to the MEC, electromethanogens can use electrons, H2, and CO2 and reduce them into CH4 (Liu et al., 2016b).

15.2.1.3.2 Inexpensive upgrading process Biogas produced from anaerobic digesters is usually a mixture of CH4 (50%
75%) and CO2 (50%
25%). Unwanted CO2 will reduce the quality of biogas, and expensive upgrading processes are required to purify it. Biogas, once purified by removing CO2, can be used as renewable and low carbon fuel for electricity generation and transportation (Yu et al., 2018). So, a reactor in which MEC is coupled with the AD is promising to produce H2 and CO2, along with H2, can be reduced to produce CH4.

15.2.1.3.3 Maintenance of reactor stability The ME process has been proved to maintain the stable operation of an AD reactor in several ways. ME process can be used to alter and control the main processes in AD. Combining AD with an ME process resulted in a higher level of biogas production and enhanced methane production. The introduction of ME in the recirculation loop of a thermophilic UASB resulted in a higher tolerance of the digester to a severe drop in pH due to the addition of an acetate pulse to the system. Insertion of the cathode electrode in an AD reactor resulted in high CH4 production and COD removal. The introduction of anode and cathode in the sludge bed of a UASB and a Continuous Stirred Tank Reactor (CSTR) also resulted in increased methane production (Liu et al., 2019). Thus ME can also be employed for postdigestion polishing of highly loaded wastewaters, leading to side products such as H2.

15.2.1.3.4 Hydrogen production increases the speed of methane production The application of various external voltages in an MEC can increase and decrease the speed and rate of hydrogen and methane production. Studies have been carried out by Cai et al. (2016a), and Guo et al. (2017) had postulated that when a voltage of 0.6
0.8 V was applied to the MEC cathode, methane composition in the biogas generation increased by 30%. Thus it is a clear fact that the ME process can increase the rate and speed of methane generation in AD.

15.2.1.4 Working principle of the reactor Biomethane is an energy product obtained biologically by two routs such as AD methanogenesis as well as through electromethanogenesis (1) via electrochemistry in which CO2 reacts with protons and electrons to be reduced to CH4 and (2) biologically by hydrogenotrophic microorganisms that improve the production rate and yield (according to Eqs. 15.i
15.iv). First, the methane is produced by the anaerobic oxidation of the wastewater along with the consumption of volatile fatty acids

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Integrated Microbial Fuel Cells for Wastewater Treatment

Figure 15.2 Schematic representation and working principle of the proposed reactor.

(VFAs) and H2 (Cai et al., 2016b; Guo et al., 2016). Then it may be produced by the reaction of electrons from the organic matter with CO2. This can be better explained as CH3 COO2 1 2H2 O ! 2CO2 1 8H1 1 8e2

(15.i)

4H1 1 8e2 ! 4H2 m

(15.ii)

CO2 1 8H1 1 8e2 ! CH4 1 2H2 O

(15.iii)

4H2 1 CO2 ! CH4 1 2H2 O

(15.iv)

A schematic representation, along with the working principle, is shown in Fig. 15.2. Integration of ME and AD reactors was done for in situ conversion of organic matter into CH4 and to treat the BW. The biogas production and organic substrate removal can be ascribed to the AD and wastewater treatment.

15.2.2 An experience of scaling up of the novel microbial electrolysis
anaerobic digestion reactor Novel integrated reactors were scaled-up for enhanced power, and organic removal performance on the basis of three major research aspects as (1) special insights into selection of cathode electrode material as electrodes play a vital role in enhancing the reactor performance, (2) determination of the influence of electrode positions

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 323

Figure 15.3 Reactor setup for the cathode-selection study. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Cui, M., Yang, C., Wang, L., et al., 2016. Cathode material as an influencing factor on beer wastewater treatment and methane production in a novel integrated upflow microbial electrolysis cell (upflow-MEC). Int. J. Hydrogen Energy 41 (4), 2189
2196.

and hydraulic retention times (HRTs) on substrate degradation and biogas production, and (3) the effects of the anode/cathode ratio were investigated in aspects of organic content removal, gas production, and current generation.

15.2.2.1 Determination of the appropriate cathode electrode material 15.2.2.1.1 Reactor construction and operation Three tubular, lab-scale, acrylic plastic, and upflow-MEC reactors (S1, S2, and S3), were designed and operated (Fig. 15.3). The working volume was 600 mL, with a height of 35 cm and a diameter of 5 cm (7:1 ratio). Granular graphite was used as the anode in all reactors. Three types of cathode mesh materials (40 woven mesh), such as stainless steel (S1), nickel (S2), and copper (S3), were used to optimize the material which could have a positive effect on methane production and organic content removal from the effluent. The distance between the electrodes was 3.0 cm, and they were connected to a data acquisition system over a 10 Ω resistance. The current was measured once in 10 min, and the applied external voltage was the 0.80 6 0.01 V. The reactors were operated at 24 h HRT, with acetate for 1 week until the current values got stabilized. Before feeding the influent was purged using high purity nitrogen gas (99.99%) for 20 min to remove oxygen. All the reactors were inoculated with 10 mL of municipal wastewater from older MECs in the lab that had been running for several months. This served as inoculum for anaerobic microbes. The reactors were run at room temperature of 30 C. Reactors were equipped with an Ag/AgCl reference electrode (RE-5B; BASi) for measuring anode potentials.

15.2.2.1.2 Sampling and electrochemical analyses Artificial BW was the influent and it was prepared in the laboratory according to 1:4 ratio of fresh beer and phosphate buffer solution (50 mM), for 1 L of influent, and in accordance with the characteristics of the real BW from Snow beer manufacturing industry in Harbin, China. Influent and effluent samples were

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Integrated Microbial Fuel Cells for Wastewater Treatment

collected for the determinations of COD, total organic carbon (TOC), soluble carbohydrates, and soluble proteins. The values were calculated along with the standard error bars. The measurement of VFAs was carried out by gas chromatography (Agilent 7890, United States), at the beginning and end of each cycle according to standard methods as previously described (Sangeetha et al., 2017). Gas was collected in gas bags (Cali-5-Bond, Calibrated Instruments Inc.) attached to the top of the reactor and it was analyzed by gas chromatography (Agilent 7890, United States) with samples taken using a gastight syringe (250 mL, Hamilton Sample-lock Syringe), and its composition was calculated according to the methods prescribed previously by Wang et al. (2009). The biogas collected in the gas bags was analyzed for CH4 (methane), H2 (hydrogen), and CO2 (carbon dioxide). The abovementioned procedures were followed for all the three research aspects carried out. Cyclic voltammetry (CV) scans were performed on plain, and biofilm covered electrodes from 0 to 21 V, at a scan rate of 0.01 V/s. The CV scans were performed to evaluate the electrochemical characteristics of the electrodes in the cathodeselection study.

15.2.2.1.3 Outcomes and substantiations The results of the study were summed up and discussed in order to identify the influence of the cathode electrode on organic content removal, biogas production, and current generation. The parameters such as COD, TOC, and carbohydrates were analyzed to calculate the organic content removal from the wastewater (Fig. 15.4), whereas methane production rate (MPR) and MY were analyzed for biogas production. S2 reactor with Ni cathode was better than SS and Cu cathodes with a maximum COD removal value of 90%, S1 documented a COD removal of

Figure 15.4 COD, carbohydrate, and TOC removal profiles. COD, Chemical oxygen demand; TOC, total organic carbon. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Cui, M., Yang, C., Wang, L., et al., 2016. Cathode material as an influencing factor on beer wastewater treatment and methane production in a novel integrated upflow microbial electrolysis cell (upflow-MEC). Int. J. Hydrogen Energy 41 (4), 2189
2196.

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 325

Figure 15.5 Methane generation with respect to cathode materials. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Cui, M., Yang, C., Wang, L., et al., 2016. Cathode material as an influencing factor on beer wastewater treatment and methane production in a novel integrated upflow microbial electrolysis cell (upflow-MEC). Int. J. Hydrogen Energy 41 (4), 2189
2196.

79%, whereas S3 had the minimum COD removal value of 69%. There was no marked difference in carbohydrate removal based on materials. But the TOC removal followed a similar trend with COD, with Ni cathode documenting a maximum of 83% removal efficiency. Cathode material greatly affected the MY of the reactors. Fig. 15.5 shows that a maximum MY of 143 mL/mg COD was observed in reactor S2, and a maximum MPR of 0.37 m3 CH4/m3 reactor/day was also measured in S2. Wang et al. (2017) reported that SS electrodes with high Ni content have the best catalytic activation and are anticorrosive, and also Ni is the most active electrode material for Hydrogen Evolution Reaction HER with the best electrocatalytic activities. So, this property might have contributed to better H2 evolution from the Ni cathode, which might have electrochemically induced high CH4 production. To evaluate the electrochemical activities and choose the best cathode material based on the lowest overpotential and highest current generation, CV was done. This was performed in a potential region of 0 to 21.0 V at a scan rate of 0.01 V/s. Fig. 15.6A shows the CV results of plain electrodes before used in the MEC. No valuable overpotential was detected in the electrodes, except the highest of 20.2 V, by Cu electrode, whereas a high current generation of 1.0 mA was detected in both SS and Ni electrodes. The used cathodes were taken from the reactors after the cycles have stopped and were analyzed along with the biofilm (Fig. 15.6B). The Cu electrode shows an overpotential of 20.22 to 20.23 V, which was much higher than the overpotential of 20.7 V, depicted by the Ni electrode. These results indicate that Ni proved to be good in the consumption of a lower voltage for proton generation and was better in current production. The overpotential values obtained in this experiment for Ni (20.7 V) was lesser than the overpotential for Pt electrodes (20.5 V).

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Integrated Microbial Fuel Cells for Wastewater Treatment

Figure 15.6 (A) Cyclic voltammogram for plain electrodes and (B) overpotential and current generation by electrodes with biofilm. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Cui, M., Yang, C., Wang, L., et al., 2016. Cathode material as an influencing factor on beer wastewater treatment and methane production in a novel integrated upflow microbial electrolysis cell (upflow-MEC). Int. J. Hydrogen Energy 41 (4), 2189
2196.

The current generation in Ni was higher at 23.3 mA than 24.0 mA in Cu. The material that generates the steepest slope for the current production at a given voltage can be selected as a cathode for reactors as it can produce more current than the others. Fig. 15.7 illustrates the current generation profile in all three reactors. The better performance of the Ni reactor is clearly indicated in the graph, with a maximum current of 8.6 mA. Stainless steel and nickel cathodes are promising in MEC due to their relatively good catalytic activity, high corrosion stability, and low cost. It was interesting to notice that Ni performed well than SS, and the reason must be the high intrinsic electrocatalytic activity of Ni than SS. Ni is deposited on SS electrodes to increase its catalytic efficiency and prevent

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 327

Figure 15.7 Current generation profile in the reactors. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Gao, L., Wang, L., Cui, M., et al., 2017. Energy recovery evaluation in an up flow microbial electrolysis coupled anaerobic digestion (ME-AD) reactor: role of electrode positions and hydraulic retention times. Appl. Energy 206, 1214
1224.

corrosion (Cai et al., 2018). Compared with stainless steel and copper meshes, nickel mesh cathode remarkably affected the reactor performance by showing promising results in organic content removal from wastewater, the current generation, and methane production. So, nickel mesh was chosen as a cathode electrode for the reactors.

15.2.2.2 Estimation of electrode positions and hydraulic retention time 15.2.2.2.1 Reactor construction and operation Electrochemically assisted methane production was studied with four lab-scale, single-chambered tubular upflow ME
AD reactors as S1, S2, S3, and S4 (Fig. 15.8). The reactor construction and the employment of optimized electrodes were according to the study of Sangeetha et al. (2016). The reactors were operated at four different HRTs as 12, 18, 24, and 36 h. The influent sample was analyzed before the start of the experiment. Effluent sampling was performed at the end of every single HRT for COD, carbohydrates, TOC, VFA, and ethanol analysis. Reactors were run in one single HRT for almost a month, and the days with steady values were selected for result representation purposes.

15.2.2.2.2 Electrochemical analyses A new method of methane generation estimation was performed as MY, MPR, HY, and hydrogen production rate (HPR) calculations. To estimate the relation between ME and AD, the methane production was calculated as current production indicated in the following equation (with respect to electron transport): Vmethane 5 UItWmethane Vm

(15.1)

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Integrated Microbial Fuel Cells for Wastewater Treatment

Figure 15.8 Experimental setup for the electrode positioning study. S1—Electrodes placed at the top of the reactor with anode placed above the cathode S2—Electrodes placed at the top of the reactor with cathode placed above the anode S3—Electrodes placed at the bottom of the reactor with anode placed above the cathode S4—Electrodes placed at the bottom of the reactor with cathode placed above the anode. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Gao, L., Wang, L., Cui, M., et al., 2017. Energy recovery evaluation in an up flow microbial electrolysis coupled anaerobic digestion (ME-AD) reactor: role of electrode positions and hydraulic retention times. Appl. Energy 206, 1214
1224.

where Vmethane is the volume of methane, U is the external voltage, I is the current, t is the time, Wmethane 5 4.043 mmol/wh indicates the conversion from electric heat into methane, and Vm 5 22.4 L/mol represents the molar volume of gas. In order to determine the relationship between ME and methane production, the methane volume was theoretically calculated with respect to the current generation considering that electrons are being donated for methane generation. It was calculated according to the equation V 5 nRT/P, where V is the volume of methane produced by electrons (mL), n is the moles of methane produced, R is the gas constant 8.314 J/mol/K, T is the temperature (308K), and P is the atmospheric pressure (105 Pa).

15.2.2.2.3 Outcomes and substantiations The results of this study identified the best position for electrode placement and optimized the appropriate HRT with more comprehensive insights into MEC reactor performance. The effects of the electrode configurations on reactor performance were investigated in aspects of organic content removal, gas production, and current generation. The results were further justified with the microbial analysis using pyrosequencing techniques. The wastewater had a COD of 2200
2250 mg/L. Fig. 15.9 shows the COD, TOC, and carbohydrate removal efficiencies recorded in the effluents of all the four reactors. Reactors with electrodes at the bottom showed better organic removal efficiencies than reactors with electrodes at the top position. An increase in HRT

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 329

Figure 15.9 Organic content removal in beer wastewater with respect to electrode positions and placement. (A) Removal efficiency of COD, Carbohydrates and TOC (B) VFA and Ethanol concentrations. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Gao, L., Wang, L., Cui, M., et al., 2017. Energy recovery evaluation in an up flow microbial electrolysis coupled anaerobic digestion (ME-AD) reactor: role of electrode positions and hydraulic retention times. Appl. Energy 206, 1214
1224.

resulted in a significant increase in organic content removal efficiencies. A maximum COD removal of 92% was documented in the effluent of the S4 reactor, whereas S1, S2, and S3 effluents showed 76%, 80%, and 88% removal efficiencies, respectively, at 36 h HRT. TOC removal also showed a similar pattern with COD removal, and S4 effluent had a better efficiency with 64%, whereas S1, S2, and S3 effluents documented 52%, 57%, and 60%, respectively, at 36 h HRT. Carbohydrates removal was also influenced by the effect of electrode configuration, where the bottom electrode-positioned reactors (S3 and S4) were better in removal (98%), and top electrodes (S1 and S2) had a lesser removal efficiency of 95%. The reason might be the increased organic oxidation in the S3 and S4 reactors, where the electrodes were arranged at the bottom might be the early and easy availability of substrate for utilization by bacteria. In the reactors where electrodes are placed at the top, the organics were made available only when the influent left the reactor as effluent. Within this time, major organics have been utilized by the suspended microbes. Placement of the anode below the cathode may also result in the effective

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Figure 15.10 Methane (A) and hydrogen (B) production in the reactors. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Gao, L., Wang, L., Cui, M., et al., 2017. Energy recovery evaluation in an up flow microbial electrolysis coupled anaerobic digestion (ME-AD) reactor: role of electrode positions and hydraulic retention times. Appl. Energy 206, 1214
1224.

transfer of protons directly to the cathode and electrons through the external circuit to the cathode for further reduction. Baek et al. (2016) operated an anaerobic digester at various HRTs, and they assured that microbial washout was of great concern while operating an AD reactor, as reduction of the microbial population could result in imbalanced reactions. Fig. 15.10A and B illustrated that electrode configuration had a substantial impact on CH4 and H2 production. Increase in HRT resulted in an increase in CH4 production, with a pattern of S4 . S3 . S2 . S1 in terms of electrode placement. Hydrogen production followed a contradictory trend to that of methane. A maximum HPR and HY of 23.8 mL H2/L reactor/day and 17.6 mL/g COD were detected in S2 at 18 h, indicating that hydrogen utilization was the least with the upper position of the cathode at the top of the reactor. In addition, low COD removal at a short HRT of 12 h reduced electron generation on the anode and consequently HPR, while long HRTs of 24 and 36 h also lead to low hydrogen collection because of hydrogen utilization by suspended methanogens. Therefore a maximum MPR of

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 331

304.5 mL CH4/L reactor/day was yielded in the S4 reactor at 36 h HRT, and then S3 with 210.3 mL CH4/L reactor/day, whereas the S1 and S2 reactors could attain a yield of 100.2 and 154.6 mL CH4/L reactor/day, respectively. The MY also followed a similar trend in the S4 reactor achieving a maximum of 275.8 mL/g COD followed by S1, S2, and S3 with a production rate of 70.9, 148.7, and 196.4 mL/g COD, respectively. Taking the abovementioned results into account, it can be inferred that the arrangement of electrodes had a remarkable effect on CH4 production. Bio electrodes at bottom position predominantly converted organics for electrons (hydrogen) production and time for hydrogen utilization in the integrated system with enhanced methane production. When the anode electrode was placed above the cathode electrode, and they were positioned at the top of the reactor, CO2 was not sufficiently able to react with the H2 produced by cathodic reduction to form CH4. It was detected that the biogas from the S1 reactor had a maximum CO2 content of 6.1%. Meanwhile, in the electrodes configured at the reactor bottom, the CO2 that was produced from bottom biofilm communities increased the chance to be well utilized by upper microbes (suspended and biofilm) to produce CH4. This may also be a reason for better CH4 generation in S3 and S4 reactors than in S1 and S2 reactors. So, these discussions well justified the better performance of the S4 reactor in both organic content removal and methane production compared to the other three reactors. The current generation is illustrated in Fig. 15.11A, S3 reactor was observed to produce a maximum current of 10 mA (0.83 mA/cm2 cathode), which was followed by S1 (9.0 mA, 0.75 mA/cm2 cathode), S4 (8.8 mA, 0.67 mA/cm2 cathode), and S2 (8.3 mA, 0.69 mA/cm2 cathode). The reason was that anode was placed above the cathode in the S3 reactor, and we assumed that the anode respiring bacteria (ARB) on the anode might have utilized the organic content as well as the H2 produced in the cathode for the synthesis of electrons. Lee et al. (2009) reported that the ARBs were utilizing H2 as the electron donor to produce electrons. It was proved that H2 is a universal electron donor for anaerobic microbes, and the ARB were likely to consume it for electron generation and also justified the fact of better performance of the S3 reactor in terms of the current generation. Therefore electrodeposition and arrangement showed an important effect on the current generation and gas production. A theoretical analysis was made to evaluate and compare the methane produced from current or electron contribution (ME methane) with the methane in the gas bag (ME
AD methane), as in Fig. 15.11B. Methane generation from electrons followed a similar pattern with the current generation, whereas methane from AD followed the trend of methane generation. It was apparent that bottom electrodepositioned reactors (S3 and S4) were the best in ME
AD methane production (189 6 18 and 218 6 19 mL, respectively) compared to the top electrode-positioned reactors (S1 and S2 with 153 6 15 and 128 6 12 mL, respectively). The current generally produced in all the four reactors at longer HRT (36 h) was lower than that produced at shorter HRT (18 h). It may be due to the fact that longer HRTs provided the substrate to the system at limiting rates, which eventually reduced the organic uptake of electrode respiring microbes. Thus eventually decreasing the electron production and thereby the current generation. Increasing the HRT from 10 to

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Figure 15.11 Current profile (A) of the reactors and methane production (B) based on electron contribution (ME) and methane measurement in accordance with electrodes configuration. ME, Microbial electrolysis. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Gao, L., Wang, L., Cui, M., et al., 2017. Energy recovery evaluation in an up flow microbial electrolysis coupled anaerobic digestion (ME-AD) reactor: role of electrode positions and hydraulic retention times. Appl. Energy 206, 1214
1224.

24 h slightly decreased the current generation, whereas decreasing it from 24 to 6 h brought about a dramatic increase in current generation activity, indicating that power generation was limited by organic supply. Upgrading of MFCs was performed by operating them at different HRTs (Pasupuleti et al., 2016). Operating continuous MFCs at short HRTs showed maximum power density due to substrate availability for anodic biofilm, resulting in higher substrate conversion efficiency and electrochemical activity than at long HRTs. But at very short HRTs, the washout of electroactive microbes occurred, resulting in a rapid fall in the current generation. The results of the pyrosequencing analysis are as follows: the major phyla, class, and genera of anode and cathode are shown in Fig. 15.12A
C. Anode biofilm had Firmicutes (41%), Proteobacteria (38%), and Bacteroidetes (16%), whereas the

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 333

Figure 15.12 Distribution of microbial community structure in the reactor biofilms of anode (A) and cathode (C) for major phyla, dominant class, and dominant genera. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Gao, L., Wang, L., Cui, M., et al., 2017. Energy recovery evaluation in an up flow microbial electrolysis coupled anaerobic digestion (ME-AD) reactor: role of electrode positions and hydraulic retention times. Appl. Energy 206, 1214
1224.

cathode biofilm had Firmicutes (46%), Proteobacteria (27%), and Bacteroidetes (21%). The six major classes such as Bacilli, Bacteroidia, Clostridia, Deltaproteobacteria, Erysipelotrichia, and Gammaproteobacteria were identified. Bacilli accounted for a maximum of 25% on the anode and 30% on cathode biofilms, where S4 anode had 40.5%, and S4 cathode had 38.6% of Bacilli on their biofilm.

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This suggested that more fermentative and oxidative communities were enriched on S4 electrodes than other electrode biofilms, suggesting enhanced organic content removal. Clostridia were enriched on S3 cathode (13.4%) and Deltaproteobacteria on S3 anode (27%), implying synergistic effects with exoelectrogens and better current production than other reactors. The class Bacteroidia consisted of abundant anaerobic bacteria responsible for decomposition and organic acid production. S1 anode and cathode had the maximum Bacteroidia of 26.3% and 22.4%, respectively, on their biofilm, which further revealed the reason behind the elevated VFA concentration in the S1 reactor than the other reactors. The class Erysipelotrichia belonged to phylum Firmicutes were recognized as a common fermenting class bacteria, which indicated that these reactors had better fermentation ability (Guo et al., 2015). Important electrogenic communities such as Geobacter and Desulfovibrio (Deltaproteobacteria and Gammaproteobacteria) were also enriched on S3 cathode biofilm, and that may be the reason for the current generation of that reactor compared to the other three reactors. Geobacter was found in fewer fractions of 3.6%, whereas Desulfovibrio was enriched with a maximum of 10.3% on S3 cathode biofilm. The most abundant genera for electron transfer were Geobacter and Desulfovibrio. Though Bacillus accounted for an average of only 2.0%
2.15% in both anode and cathode of all reactors, Lactococcus was enriched better (19% in anode and 15% in cathode) and it had developed the maximum in S4 anode with 24% and S4 cathode with 19%, suggesting a better fermentative and oxidative efficiency than the other reactors. Lactococcus also had good enrichment in S3 electrodes with 18.5% on the anode and 13.1% on the cathode. This marked a synergistic activity with electrogens (Geobacter and Desulfovibrio), and it has been remarkably reported that Lactococcus lactis possessed a self-catalyzed anodic electron transfer by excreting redox mediators (Yang et al., 2015; Tkach et al., 2017). Furthermore, there was a better prevalence of Bacteroides on the electrodes of S3 and S4 (13.4% and 14.6%, respectively), and this witnessed the better substrate degrading ability of these reactors compared with S1 and S2 reactors (5.5% and 10%, respectively). Thus it was perceptible that the key communities detected in this research were Lactococcus and Desulfovibrio belonging to classes Bacilli and Deltaproteobacteria, respectively. Methane was detected in all reactors after almost 15 days of operation. Fig. 15.13 shows the prevalence of methanogens on anode and cathode biofilm of all the reactors. Acetotrophic methanogens [Methanosarcina (facultative) and Methanosaeta (obligate)] that generate methane from acetate were predominantly found on the anode biofilm. Hydrogenotrophic methanogens (Methanobacterium, Methanobrevibacter, Methanocorpusculum, and Methanospirillum) that utilized H2 for methane production were found on the cathode biofilm, and they accounted for almost 80% of the total cathode biofilm diversity. Electrode placement inside the reactor had vital impacts on the reactor performance and functional community distribution. The influence on functional communities is depicted in Figs. 15.14 and 15.15. It can be concluded that the electrode communities were substantially determined by the electrode positions in an anaerobic system. The key functional communities were mostly methanogens, such as Methanobrevibacter, Methanosarcina,

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 335

Figure 15.13 Relative abundance of methanogens on the electrode biofilms. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Gao, L., Wang, L., Cui, M., et al., 2017. Energy recovery evaluation in an up flow microbial electrolysis coupled anaerobic digestion (ME-AD) reactor: role of electrode positions and hydraulic retention times. Appl. Energy 206, 1214
1224.

Figure 15.14 Hierarchical cluster analysis of bacterial communities from all electrode biofilms. The The Operational Taxonomic Units (OTUs) of y-axis were ordered by phylum (3% distance). Sample communities were clustered based on complete linkage method. The color intensity of scale indicates relative abundance of each OTU read. Source: Courtesy: Sangeetha, T., Guo, Z., Liu, W., Gao, L., Wang, L., Cui, M., et al., 2017. Energy recovery evaluation in an up flow microbial electrolysis coupled anaerobic digestion (ME-AD) reactor: role of electrode positions and hydraulic retention times. Appl. Energy 206, 1214
1224.

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Figure 15.15 DCA of microbial communities in electrode biofilms. DCA, Detrended correspondence analysis. Source: Courtesy: Guo, Z., Thangavel, S., Wang, L., He, Z., Cai, W., Wang, A., et al., 2016. Efficient methane production from beer wastewater in a membraneless microbial electrolysis cell with a stacked cathode: the effect of the cathode/anode ratio on bioenergy recovery. Energy Fuels 31 (1), 615
620.

Methanospirillum, Methanocorpusculum, and Methanobacterium, as well as some potential exoelectrogens, such as Geobacter and Desulfovibrio. Specifically, relatively higher anaerobic communities and exoelectrogens were detected in bioanodes when cathodes were placed below them, compared with cathodes placed above the anodes. A slight shift was demonstrated over different positions with different arrangements of bioanode and biocathode in the integrated system based on detrended correspondence analysis results (Fig. 15.15). It was revealed that fermentation bacteria, exoelectrogens, and methanogens played the primary function to complete a pathway or network in the BESs. Therefore it was worthy of evaluating the influence of electrode placement and position on the functions and performance of BESs.

15.2.2.3 Estimation of cathode/anode ratio 15.2.2.3.1 Reactor construction and operation Three single-chamber MEC reactors (R1, R2, and R3) were designed and operated in the study. They had a working volume of 700 mL and headspace of 100 mL (as seen in Fig. 15.16). The anode electrode was graphite fiber brush with a spatial volume of 78.5 cm3. The stacked cathodes were layers of circular stainless steel mesh and connected together by titanium wires. R1, R2, and R3 had two, five, and eight layers, respectively, and the ratios of cathode surface area/anode spatial volume were 1, 2.5, and 4 cm2/cm3, respectively. The electrode distance was 3 cm. External voltage was supplied by a power source (FDPS-150, Fudan Tianxin, Inc., China). Applied voltage varied from 0.5 to 0.9 V. At least, three batches (48 h for each batch) were

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Figure 15.16 Schematic representation of the reactor used for cathode/anode ratio determination. Source: Courtesy: Guo, Z., Thangavel, S., Wang, L., He, Z., Cai, W., Wang, A., et al., 2016. Efficient methane production from beer wastewater in a membraneless microbial electrolysis cell with a stacked cathode: the effect of the cathode/anode ratio on bioenergy recovery. Energy Fuels 31 (1), 615
620.

repeated for each condition to obtain stable gas production. All reactors were kept in a thermostat incubator, and the temperature was maintained at 35 6 1 C.

15.2.2.3.2 Electrochemical analysis For this study the MECs were evaluated for efficiency indexes [coulombic efficiency (ηCE), electrochemical contribution efficiency (ηECE), overall electronrecovery efficiency (ηER), and overall energy efficiency (ηEN)]. Electron balance analysis (EBA) was done for the estimation of electron transfer and utilization in the reactors. The total electrons donated by initial substrates were divided into two parts as effluent COD (Peffluent) and methane (Pmethane) as follows: Peffluent 5

Ct 3 100% C0

Ct and C0 are COD concentrations at the start and end of the experiment Pmethane 5 nCH4 3 bCH4 =C0 3 V=MO 3 bO nCH4 is the methane produced (mol), bCH4 is the moles of electrons obtained from oxidization per mole of methane (8 mol of e2/mol of methane), MO is the

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molar weight of oxygen (32 g/mol), bO is the moles of electrons transferred from organics oxidized by 1 mol of oxygen (4 mol of e2/mol of oxygen), V was the volume of the reactor (L).

15.2.2.3.3 Performance The maximum COD removal was up to 80%, but the cathode/anode ratio had no significant effects on it (Fig. 15.17A). The COD removal in R3 was 8.1% 2 8.5% higher than R1 and R2, but the methane production was influenced by the electrode ratio (Fig. 15.17B). When it increased from 1 (R1) to 4 m2/cm3 (R3), the methane production increased by 20.8%, from 0.048 to 0.058 m3/m3/day. The MPR of R3

Figure 15.17 COD removal efficiency (A) and methane generation (B) with respect to cathode/anode ratio. COD, Chemical oxygen demand. Source: Courtesy: Guo, Z., Thangavel, S., Wang, L., He, Z., Cai, W., Wang, A., et al., 2016. Efficient methane production from beer wastewater in a membraneless microbial electrolysis cell with a stacked cathode: the effect of the cathode/anode ratio on bioenergy recovery. Energy Fuels 31 (1), 615
620.

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was increased to 0.14 m3/m3/day, which increased by 1.8-fold than R1. The variation of methane recovery was calculated as the volume of methane recovered per gram of COD in an oxidized substrate and was similar to the MPR. The methane recovery in all reactors achieved 84.1
93.9 mL/g COD at the applied voltage of 0.5 V. When the voltage was 0.9 V, the methane recovery of R3 was increased to 256.7 mL/g COD, which was 1.5-fold higher than R1. Substrate utilization occurred on anodes as a result of the metabolism of anode microorganisms. Methanogenesis by electron transfer from the electrode to microbes was due to cathode. Thus the effects of cathode/anode ratio increase were significant, indicating that energy yield loss of overpotential at high applied voltages can be reduced by the application of a high cathode/anode ratio. The current generation is an indicator that the reactor has the ability of electron transfer through bioelectrochemical reactions. In this study the ratio was introduced to regulate the ability of electron generation in the anode and electron recovery in the cathode. Currents in R1, R2, and R3 at an applied voltage of 0.9 V were 12.9, 12.6, and 8.4 mA, respectively. It instructed that the bioelectrolysis process was limited by the insufficient cathode area when the cathode/anode ratio was less than 2.5 cm2/cm3. The current was not further increased when the cathode/anode ratio was above 2.5 cm2/cm3, indicating that the cathode area was not the limiting factor for improving electrochemical performance anymore. The cathodic current density in MECs significantly decreased with the enlargement of the cathode/anode ratio to 1082.1 mA/m2 in R1, to 663.5 mA/m2 in R2, and to 415.3 mA/m2 in R3. On the basis of electrochemistry principles, the overpotential of the reduction reaction that happened on the cathode (cathodic activation overpotential) has a positive correlation with the current density. Therefore applying a high surface area cathode would cause a decrease of the overpotential through a reduction of the current density and, ultimately, result in an enhancement of cathodic efficiency (Zhang and Angelidaki, 2016). The ηER and ηEN were also significantly promoted by increasing the cathode/anode ratio. The overall electron recovery in R3 was 91.2% at the voltage of 0.9 V, which was 28.6% higher than R2 (62.6%) and 56.4% higher than R1 (34.8%). It indicated that more electrons were transferred into the target product (methane) in MECs with a larger cathode/anode ratio. The overall energy-recovery efficiency was 735.1% in R3 at the voltage of 0.9 V, which was 293.1% higher than that in R2 and 356.2% higher than that in R1. Apparently, energy recovery from organic substrates was benefited by increasing the cathode area. It has been pointed out that the capital cost was an important issue for the development of BESs and technologies (Pinto et al., 2011). Considering the perspectives of operating costs and interests, maximizing the cathode/anode ratio was a feasible strategy to optimize methane production in MEC configuration design. The results of EBA (Fig. 15.18A) revealed the contributions of the bioelectrochemical reaction and sole microbial metabolism to the methane production in MECs. Under the voltage of 0.9 V, electrochemical contribution increased from 15% to 24% when the cathode/anode ratio increased twice from R1 to R2. Nevertheless, further increasing the cathode/anode ratio achieved a similar electron contribution as a result of a similar current in R2 and R3. Unlike electrochemical

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Figure 15.18 EBA analysis (A) and methane production (B) in the ME
AD reactors. AD, Anaerobic digestion; EBA, electron balance analysis; ME, microbial electrolysis.

contribution, the sole microbial contribution to methane production (Fig. 15.18B) was increasing with the increase of the cathode/anode ratio. The microbial contribution to methane in R3 (47.6% of total electrons) was much higher than that in R2 (16.4%), which explained the different methane productions between R2 and R3. Moreover, a significant decrease of the electron percentage in the effluent and the rest of the part illustrated that increasing the cathode/anode ratio could reduce the energy loss and improve wastewater treatment efficiency and overall energy conversion. As for the significant difference between experimental data and the calculated value in R1, it was probably because of the low biomass of methanogens in R1 as a result of the insufficient cathode area. Methanogens were able to obtain substrates from both the electrochemical process and AD process, when methanogen biomass was the limiting factor of methanogenesis (such as R1), and the substrates provided by the AD process might not be fully used under a closed-circuit

Integration of microbial electrolysis cells with anaerobic digestion to treat beer industry wastewater 341

condition but would be exhausted when the applied voltage was cut off. It has been proven that electrodes could provide niches for methanogens in MEC systems, but the colonization of methanogens on the anode generally had a negative effect on the bioelectrolysis performance of MEC because of the competition between methanogens and exoelectrogens (Sangeetha et al., 2017). Different from the case of the anode, the increased biomass retention on the cathode as a result of the increase of the cathode surface area maintained a stable bioelectrochemical conversion to biogas and substantially promoted the overall MY.

15.2.3 Overall summary of the experience The major summary of the experience of said three research studies to design, operate, and substantially upgrade an integrated reactor which coupled AD and MEC to accelerate methane production and simultaneously treat the beer industry wastewater. The general summary has been enlisted as follows.

15.2.3.1 Cathode selection Three types of the cathode, mesh-like stainless steel, nickel, and copper (S1, S2, and S3), respectively, were used, and the following conclusion was drawn: 1. Compared with stainless steel and copper meshes, nickel mesh cathode remarkably affected the reactor performance by showing promising results in COD with a maximum value of 84.56%, TOC removal efficiency of 83%. 2. Carbohydrate removal was not significantly affected by the cathode material, MY raised to a maximum of 0.143 mL/mg COD and MPR to 0.37 m3 CH4/m3 reactor/day in S2 reactor, with a current generation of 8.6 mA. 3. CV analysis of the cathode materials showed significant peaks, which indicated the low overpotential and high current generation of nickel. The CE values were relatively less, and we suggest that increasing the electrode size and surface area may bring about a rise in energy efficiency. 4. In-depth microbial community analysis is also needed to support the role of microbes in this study.

15.2.3.2 Electrode positions and hydraulic retention time study Studies to identify the influence of electrode position on substrate degradation and biogas production and optimize the apt HRT with more comprehensive insights into MEC reactor performance were performed. The effects of the electrode configurations on reactor performance were investigated in aspects of organic content removal, gas production, and current generation. The main conclusions of the studies are as follows: 1. Four reactors (S1, S2, S3, and S4) were constructed by placing the electrodes at different positions inside the reactor and were run at four different HRTs (48, 36, 24, and 12 h). 2. Reactors with electrodes arranged at the bottom were better than those with electrodes at the top, with S4 reactor being the best among all the reactors with maximum COD, TOC, and carbohydrate removal efficiencies of 92.1%, 64.2%, and 98.9%, respectively.

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3. MPR and MY were also highest with 304.5 mL CH4/L reactor/day and 275.8 mL/g COD, respectively, at 36 h HRT. 4. Microbial community prevalence justified the results of the study, where oxidative class Bacilli and genus Lactococcus dominated the S4 reactor biofilms and electrogens class Clostridia and Deltaproteobacteria and genus Geobacter and Desulfovibrio prevailed on the biofilms of S3 reactor.

15.2.3.3 Cathode/anode ratio study Studies were performed and determined to enhance organic content removal and improved methane generation by increasing the cathode/anode ratios in ME
AD reactors. The main conclusions of the studies are as follows: 1. Three ME
AD reactors (R1, R2, and R3) were constructed with different cathode/anode ratios of 1, 2.5, and 4 cm2/cm3, respectively. 2. Maximum current output, columbic efficiency, and cathodic current density values of 12.9 mA, 32.7%, and 664 mA/m2, respectively, were produced in the R2 reactor with 2.5 cm2/cm3 ratios. Increasing the ratio beyond that range did not yield any significant differences in output. 3. Methane conversion evaluation by EBA—almost 48% in the R3 reactor due to a large cathode surface area/anode spatial volume ratio.

Acknowledgments The authors would like to appreciate the financial support from the Ministry of Science and Technology, Taiwan, under grant number MOST 106-2218-E-027-014-MY2. The authors also acknowledge the financial support by the “Research Center of Energy Conservation for New Generation of Residential, Commercial, and Industrial Sectors” from the Featured Areas Research Center Program within the framework of the Higher Education Sprout Project by the Ministry of Education (MOE) in Taiwan.

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Integrated Microbial Fuel Cells for Wastewater Treatment

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36. Kadier, A., Simayi, Y., Kalil, M.S., Abdeshahian, P., Hamid, A.A., 2014. A review of the substrates used in microbial electrolysis cells (MECs) for producing sustainable and clean hydrogen gas. Renewable Energy 71, 466
472. Lee, H.S., Torres, C.I., Parameswaran, P., Rittmann, B.E., 2009. Fate of H2 in an upflow single-chamber microbial electrolysis cell using a metal-catalyst-free cathode. Environ. Sci. Technol. 43 (20), 7971
7976. Legras, J.L., Merdinoglu, D., Cornuet, J.M., Karst, F., 2007. Bread, beer and wine: Saccharomyces cerevisiae diversity reflects human history. Mol. Ecol. 16 (10), 2091
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339. Liu, W., He, Z., Yang, C., Zhou, A., Guo, Z., Liang, B., et al., 2016b. Microbial network for waste activated sludge cascade utilization in an integrated system of microbial electrolysis and anaerobic fermentation. Biotechnol. Biofuels 9 (1), 83. Liu, W., Wang, L., Gao, L., Wang, A.J., 2019. Hydrogen and methane production in bioelectrochemical system: biocathode structure and material upgrading. Microbial Electrochemical Technology. Elsevier, pp. 921
953. Manyuchi, M.M., Mbohwa, C., Muzenda, E., 2018. Anaerobic treatment of opaque beer wastewater with enhanced biogas recovery through Acti-zyme bio augmentation. South Afr. J. Chem. Eng. 26, 74
79. Oliveira, A.S., Baeza, J.A., Calvo, L., Alonso-Morales, N., Heras, F., Rodriguez, J.J., et al., 2019. Production of hydrogen from brewery wastewater by aqueous phase reforming with Pt/C catalysts. Appl. Catal. B: Environ. 245, 367
375. Pasupuleti, S.B., Srikanth, S., Dominguez-Benetton, X., Mohan, S.V., Pant, D., 2016. Dual gas diffusion cathode design for microbial fuel cell (MFC): optimizing the suitable mode of operation in terms of bioelectrochemical and bioelectro-kinetic evaluation. J. Chem. Technol. Biotechnol. 91 (3), 624
639. Pinto, R.P., Srinivasan, B., Escapa, A., Tartakovsky, B., 2011. Multi-population model of a microbial electrolysis cell. Environ. Sci. Technol. 45 (11), 5039
5046. Rao, A.G., Reddy, T.S.K., Prakash, S.S., et al., 2007. pH regulation of alkaline wastewater with carbon dioxide: a case study of treatment of brewery wastewater in UASB reactor coupled with absorber. Bioresour. Technol. 98, 2131
2136. Sangeetha, T., Guo, Z., Liu, W., Cui, M., Yang, C., Wang, L., et al., 2016. Cathode material as an influencing factor on beer wastewater treatment and methane production in a novel integrated upflow microbial electrolysis cell (upflow-MEC). Int. J. Hydrogen Energy 41 (4), 2189
2196. Sangeetha, T., Guo, Z., Liu, W., Gao, L., Wang, L., Cui, M., et al., 2017. Energy recovery evaluation in an up flow microbial electrolysis coupled anaerobic digestion (ME-AD) reactor: role of electrode positions and hydraulic retention times. Appl. Energy 206, 1214
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Sangeetha, T., Chen, P.T., Cheng, W.F., Yan, W.M., Huang, K.D., 2019. Optimization of the electrolyte parameters and components in zinc particle fuel cells. Energies 12 (6), 1090. Simate, G.S., Hill, A.E., 2015. 20-Water treatment and reuse in breweries. In: Brewing Microbiology. Woodhead Publishing, Oxford, United Kingdom, pp. 425
456. Simate, G.S., Cluett, J., Iyuke, S.E., Musapatika, E.T., Ndlovu, S., Walubita, L.F., et al., 2011. The treatment of brewery wastewater for reuse: state of the art. Desalination 273 (2
3), 235
247. Sinbuathong, N., Somjit, C., Leungprasert, S., 2015. Feasibility study for biohydrogen production from raw brewery wastewater. Int. J. Energy Res. 39 (13), 1769
1777. Sosa-Herna´ndez, O., Popat, S.C., Parameswaran, P., Alema´n-Nava, G.S., Torres, C.I., Buitro´n, G., et al., 2016. Application of microbial electrolysis cells to treat spent yeast from an alcoholic fermentation. Bioresour. Technol. 200, 342
349. Sun, C., Liu, F., Song, Z., Wang, J., Li, Y., Pan, Y., et al., 2019. Feasibility of dry anaerobic digestion of beer lees for methane production and biochar enhanced performance at mesophilic and thermophilic temperature. Bioresour. Technol. 276, 65
73. Tkach, O., Sangeetha, T., Maria, S., Wang, A., 2017. Performance of low temperature microbial fuel cells (MFCs) catalyzed by mixed bacterial consortia. J. Environ. Sci. 52, 284
292. Tsai, W.T., Hsu, H.C., Su, T.Y., Lin, K.Y., Lin, C.M., 2008. Removal of basic dye (methylene blue) from wastewaters utilizing beer brewery waste. J. Hazard. Mater. 154 (1
3), 73
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6448. Wang, L., Liu, W., He, Z., Guo, Z., Zhou, A., Wang, A., 2017. Cathodic hydrogen recovery and methane conversion using Pt coating 3D nickel foam instead of Pt-carbon cloth in microbial electrolysis cells. Int. J. Hydrogen Energy 42 (31), 19604
19610. Wang, C.T., Sangeetha, T., Ding, D.Q., Chong, W.T., Yan, W.M., 2018a. Implementation of surface modified carbon cloth electrodes with biochar particles in microbial fuel cells. Int. J. Green Energy 15 (13), 789
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87. Wang, C.T., Sangeetha, T., Yan, W.M., Chong, W.T., Saw, L.H., Zhao, F., et al., 2019. Application of interface material and effects of oxygen gradient on the performance of single-chamber sediment microbial fuel cells (SSMFCs). J. Environ. Sci. 75, 163
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38. Xu, F., Huang, Z., Miao, H., Ren, H., Zhao, M., Ruan, W., 2013. Identical full-scale biogaslift reactors (BLRs) with anaerobic granular sludge and residual activated sludge for brewery wastewater treatment and kinetic modeling. J. Environ. Sci. 25 (10), 2031
2040. Yang, C., Liu, W., He, Z., Thangavel, S., Wang, L., Zhou, A., et al., 2015. Freezing/thawing pretreatment coupled with biological process of thermophilic Geobacillus sp. G1: acceleration on waste activated sludge hydrolysis and acidification. Bioresour. Technol. 175, 509
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5433.

Recent advancements in scaling up microbial fuel cells

16

Soumya Pandit1, Nishit Savla2 and Sokhee P. Jung3 1 Department of Life Sciences, Sharda University, Greater Noida, India, 2Amity Institute of Biotechnology, Amity University, Mumbai, India, 3Department of Environment and Energy Engineering, Chonnam National University, Gwangju, South Korea

Chapter Outline 16.1 Introduction 349 16.2 Microbial fuel cell designs used in scale-up studies

350

16.2.1 Larger laboratory reactors 352 16.2.2 Pilot-scale tests 352

16.3 Engineering parameters affecting scale-up 16.3.1 16.3.2 16.3.3 16.3.4

355

Reactor configuration 355 Internal currents 356 Membranes 356 Tubing and compartments 356

16.4 Design limitations determined by wastewater application

357

16.4.1 Effect of buffer capacity 357 16.4.2 Influence of membrane separator 358 16.4.3 Design limitations determined by scale-up 359

16.5 Overcoming design constraints 360 16.6 Life cycle assessment 362 16.7 Current challenges and potential opportunities 16.8 Conclusion 364 References 364

16.1

363

Introduction

Clean water and energy security are considered major concerns in today’s world. Energy extraction from waste and wastewater can provide a solution for the increasing waste deposition and energy crisis. Bioenergy generation through waste treatment has instigated considerable interest and provided a new approach for the use of renewable energy sources. Therefore the production of microbial fuels has gained significant ground as a promising process for energy production and wastewater treatment (Du et al., 2007). In a microbial fuel cell (MFC), chemical energy confined in the biodegradable waste can be converted in a single step into electricity or biofuel using microorganisms. In MFC, microbes oxidize waste at the anode and

Integrated Microbial Fuel Cells for Wastewater Treatment. DOI: https://doi.org/10.1016/B978-0-12-817493-7.00016-3 © 2020 Elsevier Inc. All rights reserved.

350

Integrated Microbial Fuel Cells for Wastewater Treatment

transfer electrons to the cathode (Beyenal and Babauta, 2015). The flow of electrons can be directly utilized for power generation. MFC generates power output on the cathode via reduction. The first application of MFC was for the production of electrical current by bacteria. As new technologies developed, MFC evolved, resulting in applications for several other purposes (Jung and Regan, 2007). For the evolution of these systems, it is highly necessary to understand the exoelectrogenic bacteria and the biochemical pathways responsible for the release of electrons to the acceptors outside the cell. The main focus here is to explore various challenges and operational limitations faced during the scale-up of MFC technologies (Logan and Regan, 2006).

16.2

Microbial fuel cell designs used in scale-up studies

A typical MFC comprises three basic components, that is, an anode, a cathode, and a separator, which is commonly an ion-exchange membrane. Electrons are produced at the anode due to substrate oxidation by the action of exoelectrogenic bacteria. Such bacteria have a unique property of transferring the generated electrons to the anode’s surface. The protons generated during this process pass through the ionexchange membrane to the cathode, where reduction takes place in the presence of an electron acceptor (mostly oxygen) (Rinaldi et al., 2008). Therefore MFC performance depends on these three components. In addition, losses at any step can lead to MFC overpotential. There are several MFC designs used in laboratories. These reactor types include single- or dual-chambered cylindrical and cubic reactors, dual-chambered H cells, and tube or plate reactors. However, for most scale-up studies, tubular or flat-plate reactor designs are used (Liu et al., 2008). In scale-up studies, most configurations include a tubular anode surrounded by a separator to electrically isolate the anode from the cathode (Fig. 16.1). Typically, tubular-shaped reactors are a product of cylindrical structural materials. The supporting materials used in scaled-up reactors include polyvinyl chloride (Zhang et al., 2013a), polypropylene (Kim et al., 2009), cylindrical bottles (Kim et al., 2010), nylon tubing (Scott et al., 2007), measuring cylinders (Ga´lvez et al., 2009), and cation exchange membrane (CEM) produced into a tube. Anodes can comprise granular material, cylindrical brush, or flat electrodes made of conductive fabric and produced in a cylinder. Base materials for cathodes are carbon cloth, carbon fiber, carbon veil containing activated carbon, or platinum-based catalyst. The anode is covered by some type of membrane or separator, whereas the cathode is wrapped around the separator or membrane (Nam et al., 2017). According to most of the literature, tubular designs operating in continuous flow result in further scale-up due to the extension of the tube’s length, which derives from the connection of the additional tubular MFC modules to form an MFC stack (Kim et al., 2011). It is believed that tubular designs help to scale-up the reactor by allowing the maintenance of near-optimal cross-sectional dimensions (Scott et al., 2007).

Recent advancements in scaling up microbial fuel cells

351

Figure 16.1 Tubular brush anode reactors: (A) cylindrical bottle batch reactor, (B) perforated Polyvinyl Chloride (PVC) brush anode reactor, (C) cation exchange membrane brush anode reactor, and (D) U-shaped tubular brush anode reactor. CEM, Cation exchange membrane; R, resistor.

There are very few studies on flat-plate designs, which typically consist of regular anode chambers with a membrane or a separator material, commonly Plexiglass or any machinable plastic sandwiched between the anode and the cathode (Li et al., 2008). In these designs, anodes and cathodes are made of carbon felt, carbon paper, titanium plates, and granular graphite (the latter only for the anode). Chemical and biological catalysts can be used to catalyze the reaction at the cathode. An important design feature of flat-plates is the decreased distance between electrodes, which reduces internal resistance and increases ionic diffusion rates in comparison with other designs (Fan et al., 2012). Similar to tubular designs, flat-plate designs are also operated in a continuous flow using flat-plate modules. During scale-up, these individual flat-plate modules can be connected in series to form MFC stacks (Clauwaert et al., 2009). In both tubular and flat-plate designs, multiple MFCs can be hydraulically operated in series or parallel. In the series arrangement the influent flow sequentially goes through each MFC module, whereas in parallel, each MFC module receives the same influent (Dekker et al., 2009). Similarly, these MFC modules can be electrically connected in series or parallel to increase voltage and current, respectively.

352

Integrated Microbial Fuel Cells for Wastewater Treatment

16.2.1 Larger laboratory reactors From 1.5 µL (Qian et al., 2009) to several liters, MFC laboratory tests have been conducted with reactors that use tens of hundreds of milliliters. Considering 1 L a large-scale volume, only a few systems have been developed. They worked primarily to avoid long cycles or continuous pumping of medium volumes when larger volumes were processed (Cheng and Logan, 2011). However, it was observed that most reactors were underdesigned in terms of electrode spacing and surface area. Table 16.1 provides recent performances of MFCs with a relatively large volume ( . 1 L). When relative electrode size and spacing were preserved, the power produced by the larger fed-batch systems was consistent with the smaller ones (Li et al., 2008). The continuous-flow system, in which the liquid leaves the anode chamber and directly flows into the aerated cathode, has a simple design and low power (Li et al., 2009). Excess biomass remains in the cathode chamber unless all the biomass is consumed in the anode chamber. Data is still needed for longer operation periods. For example, a study showed that ferricyanide in stack systems was placed in a separate cathode chamber to avoid problems in the biofilm. This resulted in improved performance by changing the community of the biofilm in the anodic chamber (Aelterman et al., 2006). However, several other studies observed decreased performance with time (Zhang and Liu, 2010). Due to the differences in resistances over the stack cells or the substrate starvation in the cells during the operation (Oh and Logan, 2006), voltage reversal was also a major problem in it. This voltage reversal possibility can be minimized by avoiding low substrate concentration, which occurs in the fed-batch cycle due to the continuous flow, thereby closely matching the internal resistances in the stack among its cells. Several different MFCs were operated by the group, and each MFC was operated for 1 year at Penn State University, United States. The purpose of these reactors was to demonstrate the fan running, without any other scientific purposes. By neglecting the constant feeding the reactors could run for more than a year without apparent degradation of the fan’s performance. It was further observed that the recent MFC demonstration cell contained two cathodes, four graphite fiber brush anodes, and approximately 1 L of solution. The fan’s performance later slowed down and completely stopped after 1.5 years. The cathode was cleaned with the biofilm, and the performance was restored. These results demonstrated cathode biofouling, which was also observed to affect performance during shorter periods on laboratory systems; therefore it is the primary factor affecting the reactor’s performance (Cheng et al., 2006).

16.2.2 Pilot-scale tests MFCs were demonstrated to be useful to power remote devices for seawater applications (Cao et al., 2009). However, there are no available reports on pilot-scale tests using MFCs or microbial electrolysis cells (MECs). Nevertheless, from public conferences, online posts, and discussions with researchers it can be concluded that there are at least three pilot-scale MFCs or MECs. The first one conducted at

Table 16.1 Recent scale-up studies on microbial fuel cell. Volume

Operational mode

Substrate

Buffer

Operational temp. ( C)

Hydraulic retention time

Total operational time

COD removal (%)

Organic loading rate (g COD/l/d)

Organic removal rate (g COD/l/d)

Power density W/ m3 (W/m2)

Ref.

1

Batch

Domestic WW

100 mM

30

NA

NA

NA

NA

NA

0.33 (4.3)

B1

Continuous

Synthetic

50 mM

26

28.4 h

10 months

43

0.8

0.34

0.13 (5.6)

B1.5

Continuous

Swine WW

30

77.10

4.9

3.78

0.175 (11)

Continuous

400 days

,53

0.1 1.3

0.053 0.7

(0.37 0.31)

2.5

Batch

None

NA

NA

NA

NA

0.03

Continuous

None

Room temperature Not reported

NA

2.7

4.68 days

28 days

31

4.17

1.29

0.0018

2.7

Batch

None

Not reported

4 days

28 days

79

7.05

5.57

NA

3.6

Continuous

Primary effluent Manure slurry Landfill leachate Landfill leachate Sludge

1.21 and 4.84 days 11.1 h

NA

2

Adjusted pH increased conductivity None

Cheng and Logan (2011) Kim et al. (2011) Zhuang et al. (2012)

50 mM

35

14 days

,500 days

B60

NA

NA

0.130 (9.6)

4

Continuous

None

210 to 36

11 h

450 days

.90

0.6

0.54

NA

10

Continuous

Primary effluent Brewery

None

30

2 days

180 days

86.40

1.06

0.92

0.093 (6)

20

Continuous

None

15 weeks

60 84

0.66

0.4 0.55

0.38 (0.2)

Batch

50 mM

Room temperature 30

5 20 h

1.5

91 h

NA

88

NA

NA

1.5

Continuous

Primary effluent Sludge, fed synthetic Sludge, fed synthetic

50 mM

30

15.5 h

NA

NA

NA

NA

0.133 (2.02) 0.108

Not reported

Zhang et al. (2013a) Scott et al. (2007) Ga´lvez et al. (2009) Ga´lvez et al. (2009) Ge et al. (2013) Zhang et al. (2013b) Zhuang et al. (2012a)

Li et al. (2008) Li et al. (2008)

(Continued)

Table 16.1 (Continued) Volume

Operational mode

Substrate

Buffer

Operational temp. ( C)

Hydraulic retention time

Total operational time

COD removal (%)

Organic loading rate (g COD/l/d)

Organic removal rate (g COD/l/d)

Power density W/ m3 (W/m2)

Ref.

3.5

Continuous

Synthetic

50 mM

Not reported

5.14

.40 days

B80

0.08 0.325

0.064 0.26

0.77 (3.32)

5

Continuous

Synthetic

20 mM

30

2.15 min

37 days

NA

NA

3

2 (200)

7.5

Continuous

Synthetic

50 mM

B22

6.2 h

Several days

69 97

0.32

0.22 0.31

(2 10)

20

Continuous

Synthetic

20 mM

30

7 min

34 days

NA

NA

NA

1.44 (144)

Liang et al. (2009) Ter Heijne et al. (2011) Clauwaert et al. (2009) Dekker et al. (2009)

COD, Chemical oxygen demand; WW, wastewater.

Recent advancements in scaling up microbial fuel cells

355

Foster’s brewery in Yatala, Queensland (Australia), at the University of the Queensland by the Advanced Water Management Centre, under the direction of Jurg Keller and Korneel Rabaey (Gajda et al., 2018). The reactor has a total volume of approximately 1 m3, including 12 modules of 3 m height each, with a carbon fiber brush anode inside the tubular reactors. The upflow goes through the MFC with a graphite fiber brush cathode outside the reactor. A similar design was tested in the laboratory with ferricyanide catholyte (Rabaey et al., 2006). Little is known about MFC performance other than its limiting current generation, low solution conductivity, and excessive biochemical oxygen demand (BOD) leaving the anode chamber in the wastewater, which results in a buildup of excessive biofilm on the cathodes as wastewater is exposed to air. According to the researchers and collaborators of the University of Connecticut (Fuss and O’Neill, and HydroQual Inc.), pilot-scale MFCs are also underway in the United States (Baikun Li, personal communication). The reactor included Ptcatalyzed carbon cloth cathodes and granular graphite anodes, according to the design published by this group. These systems are essential for wastewater treatment since they can remove up to 80% of the chemical oxygen demand (COD) at concentrations of 300 600 mg/L. The first demonstration of biohydrogen production from MEC, an adaptation of MFC, was conducted at the Napa Wine Company in Oakville, CA, United States by Penn State researchers along with engineering services by Brown and Caldwell (Walnut Creek, CA, United States). The reactor design was based on stainless steel flat cathodes placed in a tank and immersing brush anodes. This reactor comprised six pairs of electrodes, with 24 modules, and approximately 1 m3 of the total volume (Das, 2018).

16.3

Engineering parameters affecting scale-up

Considering an engineering perspective, important parameters, such as reactor configuration, internal currents, membrane electrode assembly, tubing, and compartments, are discussed in the subsequent subsections.

16.3.1 Reactor configuration Reactor configuration has a major influence on MFC performance. Various designs have been developed based on efficiency and performance. For example, tubular (Kim et al., 2010), stacked, and baffled stacking tubular designs (Zhuang et al., 2012). Clauwaert et al. (2009) reported that to scale-up an MFC, reactor enlargement is not an efficient approach. The major requirement is to compartmentalize the reactor into smaller units that can be combined, thereby enabling the scale-up (Clauwaert et al., 2009). The advantages of smaller units include less internal resistance and shorter migration paths for substrate uptake and proton release, which in turn will prevent the diffusion of air near the anode, thereby preventing electrochemical losses and increasing current efficiency (CE) and power generation.

356

Integrated Microbial Fuel Cells for Wastewater Treatment

16.3.2 Internal currents Internal currents may arise due to several reasons, but most commonly due to detrimental leakage of oxygen or substrate into the anodic and cathodic compartments of the cell. These compartments are separated by a proton-exchange membrane-like, which can cause an internal short-circuit, leading to a decreased cell performance (Jung et al., 2011). A possible solution for this is to use and maintain full anaerobic conditions in the MFC. In membraneless systems, parasitic internal currents are more prominent due to higher detrimental crossover effects. Therefore the use of anaerobic conditions should be avoided. Instead, selective electrocatalysts to minimize internal currents and cross-selectivity should be preferred (Rabaey and Rozendal, 2010).

16.3.3 Membranes The MFC performance can be greatly influenced by the membrane electrode assembly. The reaction product is recovered from the cathode compartment because the membrane separates the anode and cathode chambers (Nam et al., 2017). Membranes increase electrolyte resistance and may lead to the formation of a pH gradient. Fouling of the membrane due to biofilm deposition and development of extracellular polymers, as well as replacement of proton binding sites by other cations after longterm operation, both lead to decreased conductivity and ion-exchange capacity (Scott, 2015). There are several other membranes such as CEM, anion exchange membrane (AEM), ultrafiltration membranes, bipolar membranes, and nanoporous polymer membrane (Leong et al., 2013). In bipolar membranes, additional energy is required to split the water molecule into protons and hydroxyl ions. MFC performance is better using AEM in comparison with CEM, as it lowers the transport resistance, thereby reducing the internal resistance (Roh and Woo, 2000).

16.3.4 Tubing and compartments It is also necessary to consider the use of proper gas-tight reactors, particularly for MECs, where the production of hydrogen gas is the main objective, as the hydrogen gas could leak through the long tube MFC, connectors, etc. To avoid inhibiting gas production due to hydrogen accumulation, continuous gas release methods are preferred. CE is reduced due to leakage of oxygen into the anodic chamber (Logan et al., 2006). In a previous study a biocathode based on mixed cultures of sulfurreducing bacteria was used in MFC at an applied potential of 2850 mV. The products observed were Ag/AgCl, succinate, hydrogen, glycerol, ethanol, and propionate, and this process was considered to be driven by in situ hydrogen production. However, the products changed to acetone, propionate, isopropanol, propanol, isobutyrate, isovalerate, and heptanoate, due to continuous sparging with nitrogen gas under the same operational conditions. This study suggests that electroactive biocathodes are efficient biocatalysts, and metabolic routes can be shifted by altering the headspace environment (Premier et al., 2015).

Recent advancements in scaling up microbial fuel cells

16.4

357

Design limitations determined by wastewater application

MFCs contain three main elements, based on their functionality: an anode, where oxidation occurs releasing electrons that enter the electrical circuit; a cathode, where reduction occurs, and the electrons leave the electrical circuit; and a separator, which separates anodes and cathodes to prevent electronic short-circuit while simultaneously enhancing ion transport (Mustakeem, 2015). To become an effective option for wastewater treatment, critical evaluation measures include not only economic factors, but also footprint and energy efficiency. The treatment capacity of activated sludge systems ranges from 0.5 to 2.0 kg COD/m3/ day (where m3 refers to reactor volume), while that of anaerobic systems it ranges from 8 to 20 kg COD/m3/day (Oliot et al., 2016). MFCs can reach a treatment capacity of 7 kg COD/m3/day with a density of 10 A/m2 (Rozendal et al., 2008). The main advantage of MFCs is that they manufacture energy from the organic material within the effluent. However, by doing this, MFCs directly compete with anaerobic digesters (Nancharaiah et al., 2016). Under various combinations with gas combustion engines, anaerobic digesters can manufacture electricity from organic matter of the effluent with an efficiency of 30% 35%. The energy efficiency of MFCs, however, depends on the product of coulombic and voltage efficiency. The coulombic efficiency is the fraction of electrons produced by the conversion of organic matter entering the electrical circuit. Competitive microbial processes such as methanogenesis and biomass growth reduce the coulombic efficiency since they result in less electrons/energy and, therefore, lower electricity production. Coulombic efficiency varies according to reactor design, type of wastewater, temperature, etc. It generally ranges from 5% to 38% (Wang et al., 2014). However, for synthetic wastewaters, high efficiencies, up to 100%, have been reported (Aelterman et al., 2008).

16.4.1 Effect of buffer capacity Oxidation of organic materials by electrochemically active microorganisms produces acid. Usually, for every electron, one proton is also produced. Four electrons are accepted by the molecular oxygen to reduce water. Therefore it can be calculated that 4 mmol/L H1 are produced per mmol/L COD (32 mg COD). Hence, it can be concluded that the electrical current is related to a considerable flux of protons from the bioanode to the solution. The protons must be transported to the bulk solution to forestall local acidification. For that, buffers are usually added to the anolyte and catholyte to maintain a neutral pH, which is favorable for the growth of electrochemically active microorganisms on the bioanode (Gil et al., 2003). Furthermore, the addition of a buffer can increase the solution conductivity. Usually, in laboratory-scale MFCs, a phosphate buffer of pH 7 is employed with concentrations of 10 50 mmol/L. However, concentrations as high as 320 mmol/L have been reported (Fan and Xue, 2016).

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Integrated Microbial Fuel Cells for Wastewater Treatment

Other types of the buffer can also be used. For example, Fan et al. (2007) used 0.2 mol/L of a bicarbonate buffer at pH 9 in an MFC without membrane. However, they observed a higher power output when 0.2 mol/L of phosphate buffer at pH 7 was used. It was assumed that these results were caused by the higher proton transfer rate via bicarbonate than via phosphate. Nevertheless, in batch systems such as cathodes, buffers maintain the pH neutrality for a limited period. For instance, in MFCs employing membranes, the pH within the cathode chamber eventually increases to values well over pH 12 (Rozendal et al., 2006), affecting the potential use of ion-exchange membranes. Moreover, the addition of buffers to MFC wastewater treatment is not practical due to the high prices and discharge regulations. Therefore the buffer concentration in full-scale MFCs should be restricted to the effluent pH and the bicarbonate alkalinity produced within the MFC through oxidation of organic matter by electrochemically active microorganisms (Jung and Pandit, 2019).

16.4.2 Influence of membrane separator Although MFCs without membranes are increasingly used, a significant part of MFCs still adopts membranes to separate the anode from the cathode compartment (Li et al., 2011). The charge moves through the membrane between anode and cathode to compensate for the flow of electrons outside the system. The primary MFCs, such as the Nafion 117 (DuPont Co., United States), used CEM (Rahimnejad et al., 2014). However, cations are more abundant than protons in wastewaters. Therefore cations are transported through the membrane rather than protons (Jung et al., 2011). Protons consumed within the cathode are not replenished with protons from the anode. As a result, the pH can increase within the cathode and reduce within the anode. Consequently, the system suffers a voltage loss due to the pH gradient across the membrane (Pandit et al., 2012). This voltage loss can become a major drawback, particularly in continuous full-scale systems, where there is enough time for the pH gradient to develop, and the use of high concentrations of the buffer is not economically feasible. To stop the pH rise, alternative membranes have been tested for MFC applications. These membranes include AEM, charged mosaic membranes, bipolar membranes, and ultrafiltration membranes. However, none of these membranes can satisfactorily forestall the cathode pH rise. Membranes are advantageous because they create a physical barrier between the anode and the cathode (Rahimnejad et al., 2014). Once a membrane is applied as a physical barrier, it prevents oxygen from moving from the cathode to the anode. Oxygen within the anode compartment is preferred by microorganisms as an electron acceptor over the electrode. Therefore it can decrease coulombic efficiency (Li et al., 2011). In MECs the membrane prevents the produced hydrogen from moving to the anode, where it can be consumed by methanogens (Fan and Xue, 2016). Moreover, it prevents the produced carbon dioxide in the anode chamber from mixing with the hydrogen, maintaining the hydrogen pure. Nevertheless, membranes are not 100% selective and do not forestall the mixing of anolyte and catholyte entirely. Once a flux of ions moves through the membrane, some exchange of

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anolyte and catholyte can occur (Scott, 2015). In the future, if MFCs are applied to produce special products at the cathode, this intermixture must be kept in mind. The choice to use a membrane can be based on the need to keep the products in the cathode or prevent the mixing of anolyte with catholyte.

16.4.3 Design limitations determined by scale-up 16.4.3.1 Scale-up and voltage loss The most obvious parameter that changes during scale-up is the magnitude of the current. Frequently, the current increases as a result of the increased size of a single cell and/or the number of cells electrically connected in parallel. Voltage loss can occur due to the resistance to the electron flow of the electrode material itself. However, this voltage loss is scale-dependent, as both the resistance of the electrode and the average current density decrease with size. The precise relationship between scale and voltage loss depends on the geometry of the electrode (Jung et al., 2012). Because the resistance is linearly associated with the path length of the electron flow, it can be safely assumed that the resistance increases linearly with the height of the electrode. In addition, the increase in average current through the anode is also expected to increase the height of the electrode (Kang et al., 2017). A scaled-up planar electrode was maybe a couple of meters high. A height increase with a factor of 100 can result in increased voltage loss with a factor of 10,000. In addition, the scale-up of the electrical circuitry of the MFC must also be considered, that is, whether individual cells should be electrically connected in series or parallel, which can affect voltage loss. Cells in parallel (for instance, in tubular design as tuMFC bundles) do not need to be hydraulically sealed at the end of the reactor, even if they are electrically connected. In the most straightforward configuration an MFC reactor would simply be a pile of cells with 1 cm of thickness each. These cells are often operated in parallel, that is, keeping the cells independent from one another (Pandit et al., 2017). The advantages of the parallel configuration are evident (Premier et al., 2015). First, single-cell experiments may be considered representative and prognostic for the stack performance. Second, in a properly designed stack, failure, and/or cleaning of a single-cell would not affect the performance of other cells (Liu et al., 2008). However, there are challenges in this configuration. For instance, the longdistance electrons need to travel through the electrodes. This is particularly difficult for carbon or graphite electrodes. In a scaled-up system the ohmic voltage drop could amount to several hundreds of millivolts because of the higher absolute currents in comparison with laboratory setups. Alternatively, a stack configuration with bipolar plates between the cells could be a viable design due to the shorter electron transport distances, which resulted in acceptable ohmic losses and decreased the conductivity of electrodes (Aelterman et al., 2006). Bipolar plate stack design, however, still has challenges that need to be solved, such as cell reversal (Nitisoravut et al., 2017).

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16.4.3.2 Hydrodynamics and mechanics In the future, regardless of the chosen scale-up configuration, it will be necessary to consider hydrodynamics and mechanics. A well-engineered hydrodynamic design is required to obtain and maintain an acceptable distribution of anolyte and catholyte to the cells from a shared manifold. Fluid dynamics are essential to managing the operation during pressure drop changes from possible failures of a cell. For upscale systems the cross-flow velocity and path length for fluids are essential. In a singlepass system the cross-flow velocity would become too low to produce sensible mixing, causing pH gradients (perpendicular to the biofilm). The path length would also become excessively long, causing acidification of the anodic compartment and depletion of biodegradable organic matter (in the longitudinal direction). These issues can also be encountered if serial staged stacks or recycling are applied. Energy loss because of pumping should be minimized to a certain extent. This energy loss is expected to be considerably higher than during digestion, where gas is utilized to mix the system. The mechanical strength and dimensional stability of MFC, individual cells, and its parts should also be considered. The mechanical strength of a cell and its parts during single-cell experiments is comparatively uncomplicated, as the applied areas are restricted, and dense, stiff-end plates are used to maintain the components and their shape. With increased area and number of cells in a stack the degree of distortion would increase due to the piling of (flexible) layers. Most of the topics associated with hydrodynamics and mechanics are not new and have already been explained.

16.5

Overcoming design constraints

Previously, we identified various constraints from wastewater application and scaleup. For a feasible use of MFCs for wastewater treatment, a number of constraints need to be overcome. Table 16.2 summarizes, in the initial column, the known constraints. Though this list might be discouraging, it does not imply that the MFC system will not become successful. On the contrary, neglecting these aspects can lead to failure. Overcoming these constraints is essential for a successful large-scale MFC use. Directions to resolve these constraints are given in the second column of Table 16.2. For every constraint, attainable solutions are listed. It is necessary to consider that this list of solutions addresses only the issues identified so far since the research into scale-up is on its initial stage. An MFC designed for a specific application in a specific scenario should, in fact, overcome all relevant constraints. For example, an MFC treating liquid manure has fewer limitations because of the substrate characteristics, that is, high conductivity and alkalinity. However, the resistivity of the electrode material can still be an issue, which can be negligible in small-scale experiments but crucial in largescale (Table 16.1) (Rozendal et al., 2008). Graphite, for example, is an excellent electrode material at a laboratory-scale. However, it exhibits massive ohmic losses in large-scale settings. These losses would notably emerge in large-scale single-cell

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Table 16.2 Constraints and solutions to scale-up microbial fuel cells. Constraint

Solution

Buffer capacity

High alkalinity wastewater Lower I by increasing surface area Use anode effluent as cathode influent and vice versa High COD wastewater High surface area Cheap, appropriate materials Thinner materials Production of high-value products Cheap chemical catalyst Biocathode High-conductivity wastewater Membrane electrode assembly Lower I by increasing surface area Current collector High surface area Bipolar stack designs Membraneless design More selective membranes

Reactor footprint Cost

Cathode Conductivity

High current

Membrane COD; Chemical oxygen demand.

MFC designs (i.e., the typical design of a laboratory MFC; Fig. 16.2A), as large currents need to be transported through the graphite over long distances. There are two possible solutions to overcome these ohmic losses: (1) a current collector can be used to support the electrode(s) or (2) a bipolar plate stack design (Fig. 16.2B) can be adopted. In the liquid manure scenario the primary strategy is to use a highly conductive (i.e., low resistivity) cost-efficient current collector, which is not susceptible to corrosion. This eliminates copper as a possible current collector because, despite its low price and low resistivity, it is highly susceptible to corrosion. Stainless steel, however, is not prone to corrosion, has a moderately low resistivity, and might be cost-effective. A bipolar plate stack design (i.e., the second strategy) can eliminate the need for a current collector because the travel distance for electrons will be reduced to a minimum (Fig. 16.2B). However, this design has not been thoroughly investigated for MFC applications (Shin et al., 2006). Moreover, the issue of potential cell reversal needs to be resolved before bipolar plate stack MFCs can be implemented (Oh and Logan, 2007). Regarding cost issues, it is advisable to contemplate which product, for example, electricity or hydrogen, is attractive. Hydrogen provides far higher revenue than electricity. In fact, one must consider whether hydrogen will be utilized in that particular scenario. To obtain a small footprint a high volumetric surface area is needed. If the cells are constructed with 1 cm thickness, a 100 m2/m3 surface area

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Figure 16.2 ( A) single-chambered MFC design and (B) bipolar plate stack design. MFC, Microbial fuel cell.

and a sufficiently high volumetric production would be obtained. However, it is important to ensure that the cells do not clog. Therefore a pretreatment, such as filtering, might be required. The previously given example of MFC use for the treatment of liquid manure illustrates the challenges of MFCs scale-up. Designs should be developed to overcome these challenges. The relevant constraints depend on the particular scenario and specific area unit. Most research on this topic is at a laboratory-scale, with scarce scale-up studies. Research at a laboratory-scale is required to analyze the fundamentals processes and ideas, but it cannot replace scale-up research, which is urgently required (Premier et al., 2015).

16.6

Life cycle assessment

Life cycle assessment (LCA) is a systematic set of procedures to compile and examine the inputs and outputs of energy and materials. It is a technique to determine the environmental impacts of a product, including production processes,

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usage, and disposal (Scott and Yu, 2015). This process may be costly and time-consuming. Therefore its use is limited to the analysis of produced energy over carbon emissions for various renewable technologies in the public and private sectors (Areej et al., 2014). The life cycle of a product includes the extraction of raw materials, processing/manufacturing/fabrication of the product, and transportation/distribution to consumers. Economic and engineering cost analysis must be considered while performing an LCA. Only a few LCA studies have been done on MFC. They are summarized next (Table 16.1.). Fornero et al. (2010) studied MFC in terms of electricity generation and the costs of municipal wastewater treatment (cost per 0.53 kg21 BOD, based on US standards). Considering a wastewater flow of 100 m3/day, 2 g BOD/L (assuming CE of 20% for wastewater), and a lifespan of 10 years, the net present value was calculated as US$380,528. The cost required to treat municipal wastewater with activated sludge (as a secondary treatment), which is US$32,760 per year, is significantly lower than the wastewater revenue of MFC. The economic justification for MFC was US$35,731 per year, derived from electricity generation revenue and wastewater treatment costs. This same study reported that the removal of organic matter could make the system more efficient and self-sustained compared with the generation of the electricity for MFC (not an attractive solution, considering the economic aspects) (Areej et al., 2014). Regarding energy production, industrial wastewaters are more suitable for this type of technology due to their higher concentration of contaminants and COD in comparison with municipal/domestic wastewater. Various start-up companies, such as Indian Oil Corporation Ltd. (India), Trophos Energy (United States), IntAct Labs LLC (United States), Hy-SyEnce (United States), Emefcy (Israel), and Plant-e (The Netherlands), have already developed systems based on MFC technology and are trying to commercialize them (Beyenal and Babauta, 2015). Despite the remarks and suggestions from experts and researchers, it is still too early to compare the MFC technology with other renewable energy technologies, since MFC is still under development, thereby requiring analysis of field studies and environmental conditions.

16.7

Current challenges and potential opportunities

There is a need to scale-up MFC systems and make them suitable for practical applications. The main requirement for this is to increase the size of the MFC reactor and the treatment capability to a practical level. In addition, appropriate levels of output energy should be achieved. Potential uses of MFCs include effective wastewater treatment and energy generation. However, for this technology to be ready for commercialization, various challenges need to be overcome. According to previous studies, a neutral or positive energy balance has been theoretically established. However, the operation of an energetically self-sustained MFC for wastewater treatment at field level has not been conducted (Kim et al., 2007).

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The main challenge of MFC systems for wastewater treatment is scaling up the reactor and simultaneously maintain a constant energy output per unit volume, as demonstrated by small MFC (,1 L). The most important challenge of scaling up the MFC is to increase electrochemical voltage losses (overpotential) with enlarged size. Other challenges related to the scale-up of MFC include high fabrication and material costs, operational stability issues, high internal resistance, development of other operational issues over time (Zhang and Liu, 2010), lower efficiency of mixed culture biofilm on the electrode, and slow pollutant degradation kinetics (Logan and Regan, 2006). The understanding of electrochemically active bacteria (EAB) is still in its early stages due to several unnoticed electrochemical capabilities of various microbes, which can be exploited for different MFC applications (Clauwaert et al., 2008). Due to the continuous development and research on MFC technology, several valued applications other than energy generation have been established, such as wastewater treatment and operational electronic appliances. However, to enhance power generation and quality of the treated effluent, an integrated bioprocess that combines posttreatment and phenazine methosulfate (PMS) use to improve voltage is emerging (Scott and Yu, 2015).

16.8

Conclusion

Due to the recent improvements in energy output and power generation of MFC systems, it is important to reevaluate the values and niches for practical applications (Logan, 2008). The main breakthroughs in MFC research were the use of pure culture inoculum, development of air cathode MFC, use of mediators, improvement in power management, scaling up studies, and bacterial ecology, which leads to increased power generation and understanding of kinetic challenges. Moreover, the MFC technology presents a promising area of research, which might solve problems such as energy crisis and gaseous pollution derived from fossil fuel consumption. It is necessary to further develop the MFC technology to a scalable level by using innovative design and cost-effective materials. The hybrid MFC system produced by combining wastewater treatment, resource recovery, and energy production can compensate for the costs of wastewater treatment.

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Index

Note: Page numbers followed by “f” and “t” refer to figures and tables, respectively. A Activated sludge process (ASP), 232, 235 239 AD. See Anaerobic digestion (AD) Aerobic processes, integration of BES with, 235 240, 236f Agricultural residues, 113 117, 118t Agro-industrial wastewater treatment, in MFCs, 93, 94f agricultural products processing wastewater, 109 113, 114t agricultural residues, 113 117, 118t brewery wastewater, 101 102, 101t challenges to, 120 123 dairy industry wastewater treatment, 96 100 livestock industry wastewater, 117 120, 121t molasses-based distillery wastewater, chemical physical characterization of, 112t palm oil mill effluent, 107 109, 108t, 109f, 110t starch processing wastewater, chemical physical characterization of, 111t substrate, 95 96, 99t winery wastewater, 102 106, 104t Algae-assisted cathode, in MFCs, 41, 41f Algal biocathodes, microbial fuel cells with, 218 221, 220f Algal photobioreactor, microbial fuel cells with, 215 217, 218f Anaerobic digestion (AD), 231 of beer wastewater treatment, 318 integration of BES with, 240 241, 240f microbial fuel cells coupled, 295, 298f

as biosensor for process inhibitors, 304 305 beer industry wastewater treatment, 313 nutrient recovery and toxicity removal, 303 304 treatment efficiency of, 300t undigested organics in effluent of anaerobic digestion, 299 303 Anaerobic membrane bioelectrochemical reactor (AnMBER), 256 257 Anaerobic membrane bioreactor (AnMBR), 254 Anaerobic processes, 250 251 AnMBER. See Anaerobic membrane bioelectrochemical reactor (AnMBER) AnMBR. See Anaerobic membrane bioreactor (AnMBR) Applied potential of urban water, 183 ASP. See Activated sludge process (ASP) Azo dyes, 75f, 78 79 B Beer, history of, 313 314 Beer brewery wastewater treatment, 202 203 MFC coupled AD for, 313 bioelectrochemical systems, 317 318 biological treatment methods, 316 317 brewing process and wastewater, 314 315 chemical treatment processes, 316 hydrogen production, 318 integrated MFC coupled AD, 319 342 physical treatment, 316 physicochemical parameters, 315t Benthic microbial fuel cells (BMFCs), 31 32, 33t, 40 41

370

BESs. See Bioelectrochemical systems (BESs) BET. See Bioelectrochemical treatment (BET) Beverage industry wastewater treatment, 199 beer brewery wastewater, 202 203 electricity generation, 201 substrates, 200 202 Bioanode, 56 58 Bioanode-based enhancement of dye treatment, 78 79 Biocatalysts, 138 144 Biocathode, 57 58, 59t Biocathode-based enhancement of dye treatment, 79 80 Bioeconomy, 175 176, 176f Bioelectricity, 157 158 generation, 251 252, 256 257, 260, 262 Bioelectrochemical systems (BESs), 93 94 for beer wastewater treatment, 317 318 evolution tree of, 251 252 forms of, 252, 253f history of, 4 6 integrating with wastewater treatment processes, 229 aerobic processes, 235 240, 236f anaerobic digestion, 240 241, 240f bio-electro-Fenton process, 232 235, 233f MFCs integration with dark fermentation, 242 MFCs integration with microalgae, 242 243 MFCs integration with septic tank, 241 242 membrane bioreactor coupled, 249 novel integration, 243 244 oil and petrochemical industries wastewater treatment in, 157 removal of heavy metals using, 47 approaches, 53t cathode, 57 58, 59t concept and principle, 56 57, 57f distinguished from conventional technologies, 63 66 electrode materials, 58 63, 64t stormwater treatment and energy valorization processes for, 175 two-chambered, 57f

Index

Bioelectrochemical treatment (BET), 143 Bio-electro-Fenton (BE-Fenton) process, 85, 232 235, 233f Biofuel cell, 158f Bioremediation, 164 168 Bipolar plate stack design, 362f BMFCs. See Benthic microbial fuel cells (BMFCs) Brewery wastewater (BW), 101 102 as anodic substrate in MFCs, 105t chemical physical characterization of, 101t Buffer capacity, effect of, 357 358 BW. See Brewery wastewater (BW) C CAP. See Chloramphenicol (CAP) Carbamazepine (CBZ), 143 144 Carbon dioxide, and microalgae cultivation, 222 Carbon nanotubes (CNTs), 7 Cathode of BESs, removal of heavy metals at, 57 58, 59t CBZ. See Carbamazepine (CBZ) China-global beer hub, 320 Chloramphenicol (CAP), 145 146 Chocolate industry wastewater treatment, 200 202 CNTs. See Carbon nanotubes (CNTs) COD removal, 13 14 Compartments, 356 Constructed wetlands-microbial fuel cells (CW-MFCs), 32 35, 273 basic characteristic of, 276 278, 277f challenges to, 287 288 design and operation of, 279 281, 282t future perspectives of, 287 288 integration of MFCs into, 35f merger technology, development of, 278 287 performance assessment of, 281 287, 284t probable electron transfer mechanism in, 275 276, 276f similarity with microbial fuel cell, 276 278, 279f textile wastewater treatment using, 83 84 types of, 36t Corn stover, 116

Index

Crude palm oil (CPO), 107 Current generation. See Electricity generation CW-MFCs. See Constructed wetlandsmicrobial fuel cells (CW-MFCs) D Dairy industry wastewater treatment, 96 100, 202 Dark fermentation, MFCs integration with, 242 DCF. See Diclofenac (DCF) Desalination cell-microbial fuel cells (DSMFCs), 39 40, 39f Desalination cells (DSs), 29 30 Design constraints, overcoming, 360 362, 361t, 362f Design limitations, determined by wastewater application, 357 360 buffer capacity, effect of, 357 358 hydrodynamics and mechanics, 360 membrane separator, influence of, 358 359 scale-up and voltage loss, 359 Diclofenac (DCF), 143 144 Digestate, 297 298, 303 304 Dissolved oxygen (DO) concentration and microalgae cultivation, 223 DS-MFCs. See Desalination cell-microbial fuel cells (DS-MFCs) DSs. See Desalination cells (DSs) Dye breakdown, in microbial fuel cells, 74 76, 75f microbial diversity, 81 82 Dye removal, in microbial fuel cells, 76 77 E EABs. See Electrochemically active bacteria (EABs) EFCs. See Enzymatic fuel cells (EFCs) Electricity generation, 29 30 algae-assisted cathode in MFCs, 41 constructed wetlands-microbial fuel cells, 32 35, 36t MBR-microbial fuel cells, 35 39 in microbial fuel cells, 76 77 sediment microbial fuel cells, 31 32, 33t shuttle effect on, 81

371

Electrochemically active bacteria (EABs), 93 94 Electrocoagulation, 163 Electroconductivity of urban water, 181 Electrode materials, used for heavy metal removal in BESs, 58 63, 64t Electrodes of urban water, 183 Electro-Fenton process, integration of BES with. See Bio-electro-Fenton (BEFenton) process Electro-Fenton reactions, 40 Electrolysis cell with microbial fuel cell (MFC-MEC), 85 86 Electron transfer mechanism of urban water, 182 183 Emerging contaminants, control of, 262 267 Energy valorization processes, bioelectrochemical systems for, 175 challenges to, 189 191 future perspectives of, 189 191, 190f environmental economics, 185 186 environmental impact of, 186 187 influencing factors, 179 185 electroconductivity, 181 electrodes and applied potential, 183 electron transfer mechanism, 182 183 kinetics and thermodynamics, 182 membranes, 183 185 pH, 180 181 temperature, 181 life cycle analysis, 186 187 outlook, 189 191 technical scales, 187 189, 189f Environmental economics, 185 186 Environmental impact of stormwater treatment, 186 187 Enzymatic fuel cells (EFCs), 157 158, 168 Evolution tree of bioelectrochemical systems, 251 252 F Fenton oxidation, 17 Field-scale applications, 267 Floating macrophytes-based microbial fuel cells (FMFCs), 31 32, 33t Flocculation, 163 Flotation, 163

372

FMFCs. See Floating macrophytes-based microbial fuel cells (FMFCs) Food processing wastewater treatment, 199 electricity generation, 201 substrates, 200 202 Free water surface (FWS) wetlands, 273 274 FWS. See Free water surface (FWS) wetlands G GAC-based MFC, dye removal from real dye wastewater in, 76 77 GAC-based single-chamber MFC (GACBSCMFC), 79 80, 82 83 GACB-SCMFC. See GAC-based singlechamber MFC (GACB-SCMFC) H Heavy metals pollutants, 51t removal of, using BESs, 47 advantages, 66f approaches, 53t cathode, 57 58, 59t concept and principle, 56 57, 57f distinguished from conventional technologies, 63 66 electrode materials, 58 63, 64t Horizontal subsurface flow (HSSF) wetlands, 273 274, 279 281 HRT. See Hydraulic retention time (HRT) HSSF. See Horizontal subsurface flow (HSSF) wetlands Hybrid BES MBR system, 254 262, 255f combined BES MBR system, 259 262 membrane bioreactor as post-treatment unit, 260 262 membrane bioreactor as pretreatment unit, 260 integrated BES MBR system, 255 259 membrane as anode-cum-filtration unit, 258 membrane as cathode-cum-filtration unit, 255 258 membrane as separator-cum-filtration unit, 259, 259f performance evaluation of, 263t

Index

Hydraulic retention time (HRT), 327 336, 329f, 330f, 332f, 333f, 335f, 336f, 341 342 Hydrodynamics, 360 Hydrogen production, in anaerobic reactors with beer wastewater, 318, 321 I Ibuprofen (IBF), 143 144 Integrated MFC coupled AD, 319 342 application prospects, 320 cathode electrode material, determination of, 322f, 323 327, 323f, 324f, 325f, 326f, 327f cathode selection, 341 cathode/anode ratio, estimation of, 336 342, 337f, 338f, 340f China-global beer hub, 320 complete treatment, 320 321 electrode positions, estimation of, 327 336, 328f, 341 342 hydraulic retention time, 327 336, 329f, 330f, 332f, 333f, 335f, 336f, 341 342 hydrogen production and speed of methane production, 321 inexpensive upgrading process, 321 reactor stability, maintenance of, 321 scaling up, 322 341 significance of, 319f, 320 working principle, 321 322 Integrated technologies, 29 Intensified constructed wetlands, 275 Internal currents, 356 Italian Environmental Agency (APAT), 96 97 K Kinetics of urban water, 182 L Lacunas, 250 251 Larger laboratory reactors, 352, 353t Life cycle analysis (LCA), 362 363 of stormwater treatment, 186 187 Light intensity, and microalgae cultivation, 221 222 Livestock industry wastewater, 117 120, 121t

Index

M MBR-MFCs. See MBR-microbial fuel cells (MBR-MFCs) MBR-microbial fuel cells (MBR-MFCs), 35 39 general design of, 38f MBRs. See Membrane bioreactors (MBRs) MDCs. See Microbial desalination cells (MDCs) Mechanics, 360 MECs. See Microbial electrolysis cells (MECs) Membrane-based enhancement of dye treatment, 80 81 Membrane biofilters (MBRs), 29 30 Membrane bioreactor coupled bioelectrochemical systems, 249 future perspectives of, 262 267 emerging contaminants, control of, 262 267 field-scale applications, 267 membrane fouling mitigation, 262 water energy nexus, 262 hybrid BES MBR system. See Hybrid BES MBR system Membrane bioreactors (MBRs), 252 254 Membrane fouling mitigation, 262 Membranes, 356 proton-exchange, 6 7, 12, 74 75, 77, 80, 184 separator, influence of, 358 359 of urban water, 183 185 MFC-ABOR. See MFC-coupled aerobic biocontact oxidation reactor (MFCABOR) system MFC-coupled aerobic biocontact oxidation reactor (MFC-ABOR) system, 84 MFCs. See Microbial fuel cells (MFCs) MFM. See Microfiltration membrane (MFM) Microalgae, MFCs integration with, 242 243 Microalgae cultivation, for wastewater treatment and energy recovery, 211 algal biocathodes, 218 221, 220f algal photobioreactor, 215 217, 218f factors influencing, 221 223 carbon dioxide, 222 dissolved oxygen concentration, 223 light intensity, 221 222

373

pH, 222 performance of, 216t Microbial desalination cells (MDCs), 5, 5f Microbial diversity, involvement in dye breakdown in MFCs, 81 82 Microbial electrolysis cells (MECs), 4 5, 5f, 56 57, 137f, 252 Microbial fuel cell-coupled techniques, for textile wastewater treatment, 83 86 bioelectro-Fenton technology-microbial fuel cell, 85 electrolysis cell with microbial fuel cell (MFC-MEC), 85 86 MFC-coupled aerobic biocontact oxidation reactor system, 84 microbial fuel cell-integrated constructed wetlands, 83 84 Microbial fuel cell-integrated wastewater treatment systems, 29 constructed wetlands-microbial fuel cells, 32 35, 35f, 36t desalination cell-microbial fuel cells, 39 40, 39f MBR-microbial fuel cells, 35 39, 38f other processes, 40 41 sediment microbial fuel cells, 30 32, 31f, 33t Microbial fuel cells (MFCs), 1, 5f, 136f, 158f, 213 214 agro-industrial wastewater treatment in, 93 beverage industry wastewater treatment in, 199 bipolar plate stack design, 362f challenges to, 6 14 electrode materials, characteristics of, 8t food processing wastewater treatment in, 199 future of, 14 17 history of, 4 6 for industrial wastewater treatment, performance of, 15t pharmaceutical wastewater treatment in, 135 principles of, 6 14 single-chamber, 12, 13f, 200, 362f scale-up studies, 347 textile wastewater treatment using, 73 two-chamber, 12, 13f, 199 200

374

Microfiltration membrane (MFM), 80 Micro-flocculation, 163 MLE. See Modified Ludzack Ettinger (MLE) process Modified Ludzack Ettinger (MLE) process, 237 238 Molasses-based distillery wastewater, 111 112 chemical physical characterization of, 112t MTWW. See Mustard tuber wastewater (MTWW) Mustard tuber wastewater (MTWW), 111 N Nafion, 6 7 Nafion 117 membrane, 231 NASA. See National Aeronautics and Space Administration (NASA) National Aeronautics and Space Administration (NASA), 4 90 L stackable baffled microbial fuel cell, 103f Nutrient recovery, using MFC coupled anaerobic digestion, 303 304 O Oil and petrochemical industries wastewater treatment, in BECs, 157 advantages and disadvantages of, 159t conventional treatment process, 158 164, 161f, 162f research attempts, 165t substrates, 167t Olive washing wastewater (OWW), 112 113 Orange G (OG) dye, 77 Organic loading rates (OLRs), 229, 233 234 ORR. See Oxygen reduction reaction (ORR) OWW. See Olive washing wastewater (OWW) Oxygen reduction reaction (ORR), 215 217 P Palm oil mill effluent (POME), 107 109 chemical physical characterization of, 108t as substrate in MFCs, 110t

Index

wastewater discharge of, 109f Paracetamol (PAM), 146, 234 235 PEM. See Proton-exchange membrane (PEM) pH and microalgae cultivation, 222 of urban water, 180 181 PhACs. See Pharmaceutically active compounds (PhACs) Pharmaceutically active compounds (PhACs), 138 145 Pharmaceutical wastewater (PW) treatment, in MFCs, 135 applications of, 138 146, 139t composition, 142t paracetamol, 146 potential and challenges, 147 148 SBR-BET, 146 stacked constructed wetlands, 147 Photosynthetic microbial fuel cells, 252 Pilot-scale tests, 352 355 Plant microbial fuel cells (PMFCs), 32 PMFCs. See Plant microbial fuel cells (PMFCs) Polishing unit, 298 POME. See Palm oil mill effluent (POME) Probable electron transfer mechanism, in CW-MFCs, 275 276, 276f Proton-exchange membrane (PEM), 6 7, 12, 74 75, 77, 80, 184 R Reactor configuration, 355 Recalcitrant, 143 Redox gradient of CW-MFCs, 276 281 Rice straw, 116 S SBRs. See Sequential batch reactors (SBRs) Scale-up, 40 Scale-up studies on microbial fuel cells, 40, 350 355 current challenges and potential opportunities, 363 364 design constraints, overcoming, 360 362, 361t, 362f design limitations, determined by wastewater application, 357 360 buffer capacity, effect of, 357 358

Index

hydrodynamics and mechanics, 360 membrane separator, influence of, 358 359 scale-up and voltage loss, 359 engineering parameters affecting, 355 356 internal currents, 356 membranes, 356 reactor configuration, 355 tubing and compartments, 356 larger laboratory reactors, 352, 353t life cycle assessment, 362 363 pilot-scale tests, 352 355 recent advancements in, 347 Sediment microbial fuel cells (SMFCs), 30 32, 252 forms of, 33t general design of, 31f Septic tank, MFCs integration with, 241 242 Sequential batch reactors (SBRs), 146 Shuttle effect, on dye removal and electricity generation, 81 Slaughterhouse wastewater, chemical physical characterization of, 119t SL-MFCs. See Soil-microbial fuel cells (SLMFCs) SMFCs. See Sediment microbial fuel cells (SMFCs) SMX. See Sulfamethoxazole (SMX) Soil-microbial fuel cells (SL-MFCs), 31 32, 33t Stacked constructed wetlands, 147 Starch processing wastewater (SPW), 111 112 chemical physical characterization of, 111t Stormwater treatment, bioelectrochemical systems for, 175 challenges to, 189 191 environmental economics, 185 186 environmental impact of, 186 187 future perspectives of, 189 191, 190f influencing factors, 179 185 electroconductivity, 181 electrodes and applied potential, 183 electron transfer mechanism, 182 183 kinetics and thermodynamics, 182

375

membranes, 183 185 pH, 180 181 temperature, 181 life cycle analysis, 186 187 outlook, 189 191 technical scales, 187 189, 189f urban stormwater, 177 179 pollutants, characterization of, 178t quality runoff, 180t Sulfamethoxazole (SMX), 143 Sulfonamides, 144 T TEA. See Terminal electron acceptor (TEA) Temperature of urban water, 181 Terminal electron acceptor (TEA), 93 94 Textile wastewater treatment, using MFCs, 73 current generation, 76 77 dye breakdown, 74 76, 75f microbial diversity, 81 82 dye removal, 76 77 MFC-coupled techniques, for textile wastewater treatment, 83 86 bioelectro-Fenton technology-microbial fuel cell, 85 electrolysis cell with microbial fuel cell (MFC-MEC), 85 86 MFC-coupled aerobic biocontact oxidation reactor system, 84 microbial fuel cell-integrated constructed wetlands, 83 84 performance enhancement, 78 81 bioanode-based enhancement, 78 79 biocathode-based enhancement, 79 80 membrane-based enhancement of dye treatment, 80 81 shuttle effect on dye removal and electricity generation, 81 research gap, 86 total COD removal, 77 toxicity, 82 83 Thermodynamics of urban water, 182 TMP. See Transmembrane pressure (TMP) Total COD removal, 77 Toxicity removal, using MFC coupled anaerobic digestion, 303 304 TPTC. See Triphenyltin chloride (TPTC) Transmembrane pressure (TMP), 253 254 Triphenyltin chloride (TPTC), 235

376

Tubing, 356 Tubular brush anode reactors, 350, 351f U UASB. See Upflow anaerobic sludge blanket (UASB) reactor Undigested organics in effluent of anaerobic digestion, MFC coupling for treatment to, 299 303 Upflow anaerobic sludge blanket (UASB) reactor, 229, 231 232, 297 298 Urban stormwater, 177 179 pollutants, characterization of, 178t quality runoff, 180t V Voltage loss, Scale-up and, 359 W Wastewater, sources of, 249 250

Index

Wastewater treatment See also individual wastewater treatments conventional practices of, 250 251 systems, MFC-integrated, 29 constructed wetlands-microbial fuel cells, 32 35, 35f, 36t desalination cell-microbial fuel cells, 39 40, 39f MBR-microbial fuel cells, 35 39, 38f other processes, 40 41 sediment microbial fuel cells, 30 32, 31f, 33t Water energy nexus, 262 Wet oxidation, 163 164 Wheat straw, 113 116 Winery wastewater (WW), 102 106 chemical physical characterization of, 104t WW. See Winery wastewater (WW)