Delivering Low-Carbon Biofuels with Bioproduct Recovery: An Integrated Approach to Commercializing Bioelectrochemical Systems 9780128218419, 012821841X

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Delivering Low-Carbon Biofuels with Bioproduct Recovery

Bioelectrochemical Systems: The Way Forward – Volume I

Delivering Low-Carbon Biofuels with Bioproduct Recovery An Integrated Approach to Commercializing Bioelectrochemical Systems Edited by

Lakhveer Singh Department of Environmental Science, SRM University-AP, Amaravati, India

Durga Madhab Mahapatra TERI Deakin Nanobiotechnology Center (TDNBC), TERI, India

Elsevier Radarweg 29, PO Box 211, 1000 AE Amsterdam, Netherlands The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, United Kingdom 50 Hampshire Street, 5th Floor, Cambridge, MA 02139, United States Copyright © 2021 Elsevier Inc. All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Details on how to seek permission, further information about the Publisher’s permissions policies and our arrangements with organizations such as the Copyright Clearance Center and the Copyright Licensing Agency, can be found at our website: www.elsevier.com/permissions. This book and the individual contributions contained in it are protected under copyright by the Publisher (other than as may be noted herein). Notices Knowledge and best practice in this field are constantly changing. As new research and experience broaden our understanding, changes in research methods, professional practices, or medical treatment may become necessary. Practitioners and researchers must always rely on their own experience and knowledge in evaluating and using any information, methods, compounds, or experiments described herein. In using such information or methods they should be mindful of their own safety and the safety of others, including parties for whom they have a professional responsibility. To the fullest extent of the law, neither the Publisher nor the authors, contributors, or editors, assume any liability for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions, or ideas contained in the material herein. British Library Cataloguing-in-Publication Data A catalogue record for this book is available from the British Library Library of Congress Cataloging-in-Publication Data A catalog record for this book is available from the Library of Congress ISBN: 978-0-12-821841-9 For Information on all Elsevier publications visit our website at https://www.elsevier.com/books-and-journals

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Contents List of Contributors ..................................................................................................xi

CHAPTER 1 Electrical energy produced by microbial fuel cells using wastewater to power a network of smart sensors .......................................................................... 1 1.1 1.2

1.3 1.4 1.5

P.M.D. Serra and A. Esp´ırito-Santo Introduction ....................................................................................1 Microbial fuel cells ........................................................................4 1.2.1 Microbial fuel cells theoretical analysis............................. 4 1.2.2 Energy extraction from microbial fuel cells ...................... 6 Energy production, regulation and storage..................................21 1.3.1 Energy regulation and storage .......................................... 21 Smart sensor structure and operation...........................................25 Conclusions ..................................................................................27 Acknowledgments ....................................................................... 27 References.................................................................................... 28

CHAPTER 2 Application of bioelectrochemical systems in wastewater treatment and hydrogen production ....... 31

2.1 2.2 2.3 2.4 2.5 2.6 2.7 2.8 2.9 2.10 2.11

Santhana Krishnan, Abudukeremu Kadier, Mohd Fadhil Bin MD Din, Mohd Nasrullah, Nurul Nazleatul Najiha, Shazwin Mat Taib, Zularisam Ab Wahid, Yu You Li, Yu Qin, Kamal Kishore Pant, Shreesivadasan Chelliapan, Hesam Kamyab, Imran Ahmad and Lakhveer Singh Introduction ..................................................................................31 MEC fundamentals and working principles ................................32 Electron transfer mechanism........................................................34 MEC technology in hydrogen production using wastewater ......35 Agro wastewater...........................................................................36 Domestic waste water ..................................................................36 Industrial wastewater....................................................................38 Fermentation effluent ...................................................................38 Nutrient and heavy metals removals in MEC .............................39 Integrated MEC approach ............................................................40 Conclusions ..................................................................................42 Acknowledgments ....................................................................... 42 References.................................................................................... 42

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CHAPTER 3 Nutrient removal and recovery in bioelectrochemical systems....................................... 45 Aryama Raychaudhuri and Manaswini Behera 3.1 Introduction ..................................................................................45 3.2 Nitrogen removal and recovery ...................................................47 3.2.1 Issues related to conventional technologies ..................... 48 3.2.2 Nitrogen removal in bioelectrochemical system.............. 49 3.2.3 Ammonia recovery............................................................ 59 3.2.4 Challenges in nitrogen removal and recovery.................. 62 3.3 Phosphorus removal and recovery ...............................................66 3.3.1 Issues related to biological phosphorus removal ............. 66 3.3.2 Struvite precipitation......................................................... 68 3.3.3 Phosphorus removal and recovery in bioelectrochemical system ................................................ 69 3.3.4 Challenges in phosphorus removal and recovery............. 71 3.4 Conclusion and future perspectives .............................................74 References.................................................................................... 75

CHAPTER 4 Role of bioelectrochemical systems for bioremediation of wastewaters and bioenergy production ................................................................... 85

4.1 4.2 4.3

4.4

4.5 4.6 4.7

Muhammad Faisal Siddiqui, Zahid Ullah, Lakhveer Singh, Farhana Maqbool, Sadia Qayyum, Ihsan Ullah, Ziaur Rahman and Fazal Adnan Introduction ..................................................................................85 Principle of bioelectrochemical systems .....................................86 Kinds of bioelectrochemical systems ..........................................87 4.3.1 Microbial fuel cells ........................................................... 87 4.3.2 Microbial electrolysis cells for energy ............................. 88 4.3.3 Microbial electrosynthesis for energy production............ 88 4.3.4 Enzymatic fuel cells for energy production ..................... 88 4.3.5 Microbial solar cells for energy production ..................... 89 4.3.6 Plant microbial fuel cells for energy production ............. 89 4.3.7 Microbial desalination cells for energy production ......... 90 Role of bioelectrochemical systems in remediation of pollutants ......................................................................................90 4.4.1 Remediation of organic xenobiotics ................................. 90 4.4.2 Treatment of inorganic pollutants .................................... 92 Sustainability of the technology ..................................................93 Scaling up of the technology .......................................................94 Conclusion ....................................................................................94

Contents

Acknowledgments ....................................................................... 95 References.................................................................................... 95

CHAPTER 5 Energy generation from fish-processing waste using microbial fuel cells ........................................ 101

5.1 5.2

5.3

5.4

5.5

5.6

A.R. Abdul Syukor, Suryati Sulaiman, Jadhav Pramod Chandrakant, Puranjan Mishra, Mohd Nasrullah, Lakhveer Singh and A.W. Zularism Introduction ................................................................................101 National Green Technology Policy............................................103 5.2.1 Waste from fresh markets............................................... 103 5.2.2 Fish-processing wastewater characteristics .................... 105 Microbial fuel cell system..........................................................107 5.3.1 Substrates used in microbial fuel cell............................. 109 5.3.2 Fish-processing waste as substrate ................................. 109 Treatment methodology of fish-waste using microbial fuel cell (a Malaysian case study).....................................................110 5.4.1 Preparing the substrate.................................................... 110 5.4.2 Testing for physical, chemical, and biological parameters ....................................................................... 110 5.4.3 Electrode ......................................................................... 111 Results observation ....................................................................113 5.5.1 Voltage production.......................................................... 113 5.5.2 Biochemical oxygen demand removal ........................... 115 5.5.3 Chemical oxygen demand removal ................................ 115 5.5.4 Nitrogen........................................................................... 116 5.5.5 Phosphorous .................................................................... 118 Conclusion ..................................................................................119 References.................................................................................. 120

CHAPTER 6 Microbial electrosynthesis: Recovery of high-value volatile fatty acids from CO2 .................. 123 Narnepati Krishna Chaitanya, Akash Tripathi and Pritha Chatterjee 6.1 Introduction ................................................................................123 6.2 Basic principle of microbial electrosynthesis cell.....................124 6.3 Factors affecting product titer....................................................124 6.3.1 The effect of pH.............................................................. 125 6.3.2 Fluctuations in electricity supply.................................... 126 6.3.3 Impact of inoculum......................................................... 127

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6.4 6.5 6.6 6.7

6.3.4 Electrode materials.......................................................... 128 6.3.5 Effect of electrode potential ........................................... 129 6.3.6 Effect of reactor design................................................... 130 Strategies to improve product titer ............................................131 Economic evaluation ..................................................................134 Future scope of work .................................................................136 Conclusion ..................................................................................137 References.................................................................................. 137

CHAPTER 7 Low carbon fuels and electro-biocommodities ....... 143 7.1 7.2 7.3 7.4

7.5 7.6

Bahaa Hemdan, S. Bhuvanesh and Surajbhan Sevda Introduction ................................................................................143 Working mechanism of bioelectrochemical systems ................144 Application of microbial electrochemical technologies in wastewater treatment..................................................................146 Electro-biocommodities and value-added biochemical’s production...................................................................................148 7.4.1 Biohydrogen production ................................................. 149 7.4.2 Biomethane production ................................................... 149 7.4.3 Bioethanol production..................................................... 151 7.4.4 Acetate production .......................................................... 152 7.4.5 Hydrogen peroxide production ....................................... 153 7.4.6 Other value-added biochemical production.................... 153 Recent progress for electro-biocommodities generation in a bioelectrochemical system .........................................................154 Conclusion ..................................................................................156 Acknowledgment ....................................................................... 156 References.................................................................................. 157

CHAPTER 8 Potential of high energy compounds: Biohythane production ................................................................. 165 8.1 8.2 8.3 8.4 8.5 8.6

Surajbhan Sevda, Vijay Kumar Garlapati, Swati Sharma and T.R. Sreekrishnan Introduction ................................................................................165 Main aspects of the biohythane generation in bioelectrochemical system .........................................................166 Substrate for biohythane generation ..........................................167 Recent progress for biohythane generation in bioelectrochemical system .........................................................169 Use of biohythane ......................................................................172 Future prospects and concluding remarks .................................173

Contents

Acknowledgment ....................................................................... 173 References.................................................................................. 173

CHAPTER 9 Biological and chemical remediation of treated wood residues ........................................................... 177 9.1 9.2 9.3

9.4

Lais Gonc¸alves da Costa, Yonny Martinez Lopez, Victor Fassina Brocco and Juarez Benigno Paes Introduction ................................................................................177 Environmental risks of treated wood.........................................178 Remediation and recovery of treated wood...............................180 9.3.1 Bioremediation ................................................................ 180 9.3.2 Mechanisms used by fungi in the remediation process ............................................................................. 183 9.3.3 Chemical remediation ..................................................... 186 Concluding remarks ...................................................................189 References.................................................................................. 189

CHAPTER 10 An overview on degradation kinetics of organic dyes by photocatalysis using nanostructured electrocatalyst .......................................................... 195 Rishu Katwal, Richa Kothari and Deepak Pathania Introduction ................................................................................195 Organic dyes...............................................................................196 Classification of organic dyes....................................................196 Methods for the removal of pollutants ......................................196 Advanced oxidation processes ...................................................197 Photocatalysis .............................................................................198 Photocatalysts .............................................................................200 Photocatalyst surface modifications ..........................................201 Kinetics of photocatalytic degradation ......................................202 Photocatalytic reaction parameters ............................................203 Photocatalytic activity of nonmetals and metalloids supported nanophotocatalyst ......................................................204 10.12 Photocatalytic activity of polymer supported nanophotocatalyst .......................................................................206 10.13 Conclusions ................................................................................207 References.................................................................................. 207 10.1 10.2 10.3 10.4 10.5 10.6 10.7 10.8 10.9 10.10 10.11

Index ......................................................................................................................215

ix

List of Contributors A.R. Abdul Syukor Faculty of Civil Engineering Technology, Universiti Malaysia Pahang (UMP), Kuantan, Malaysia Fazal Adnan Atta ur Rahman School of Applied Biosciences, National University of Sciences & Technology, Pakistan Imran Ahmad Department of Engineering, Razak Faculty of Technology and Informatics, Universiti Teknologi Malaysia, Jalan Sultan Yahya Petra, Kuala Lumpur, Malaysia Manaswini Behera School of Infrastructure, Indian Institute of Technology Bhubaneswar, Bhubaneswar, India S. Bhuvanesh Director’s Research Cell, CSIR-National Environmental Engineering Research Institute, Nagpur, India Victor Fassina Brocco Center for Higher Studies of Itacoatiara, Amazonas State University (CESIT/ UEA), Itacoatiara, Brazil Narnepati Krishna Chaitanya Department of Civil Engineering, Indian Institute of Technology Hyderabad, Hyderabad, India Jadhav Pramod Chandrakant Faculty of Civil Engineering Technology, Universiti Malaysia Pahang (UMP), Kuantan, Malaysia Pritha Chatterjee Department of Civil Engineering, Indian Institute of Technology Hyderabad, Hyderabad, India Shreesivadasan Chelliapan Department of Engineering, Razak Faculty of Technology and Informatics, Universiti Teknologi Malaysia, Jalan Sultan Yahya Petra, Kuala Lumpur, Malaysia Lais Gonc¸alves da Costa Department of Forest and Wood Science, Federal University of Esp´ırito Santo, Jeroˆnimo Monteiro, Brazil A. Esp´ırito-Santo Department of Electromechanical Engineering, University of Beira Interior, Covilha˜, Portugal; IT—Institute of Telecommunications, Covilha˜, Portugal

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Mohd Fadhil Bin MD Din Centre for Environmental Sustainability and Water Security (IPASA), Research Institute of Sustainable Environment (RISE), School of Civil Engineering, Faculty of Engineering, Universiti Teknologi Malaysia, Skudai, Malaysia Vijay Kumar Garlapati Department of Biotechnology and Bioinformatics, Jaypee University of Information Technology, Waknaghat, India Bahaa Hemdan Department of Biosciences and Bioengineering, Indian Institute of Technology Guwahati, Guwahati, India; Water Pollution Research Department, Environmental Research Division, National Research Centre, Giza, Egypt Abudukeremu Kadier Laboratory of Environmental Science and Technology, The Xinjiang Technical Institute of Physics and Chemistry, Key Laboratory of Functional Materials and Devices for Special Environments, Chinese Academy of Sciences, Urumqi, China Hesam Kamyab Department of Engineering, Razak Faculty of Technology and Informatics, Universiti Teknologi Malaysia, Jalan Sultan Yahya Petra, Kuala Lumpur, Malaysia Rishu Katwal Department of chemistry, CSKHPKV, Palampur, India Richa Kothari Department of Environmental Sciences, Central University of Jammu, Bagla (Rahya-Suchani), Samba, Jammu & Kashmir, India Santhana Krishnan Centre for Environmental Sustainability and Water Security (IPASA), Research Institute of Sustainable Environment (RISE), School of Civil Engineering, Faculty of Engineering, Universiti Teknologi Malaysia, Skudai, Malaysia Yu You Li Department of Civil and Environmental Engineering, Graduate School of Engineering, Tohoku University, Sendai, Japan Yonny Martinez Lopez Department of Forest and Wood Science, Federal University of Esp´ırito Santo, Jeroˆnimo Monteiro, Brazil Farhana Maqbool Department of Microbiology, Hazara University, Mansehra, Pakistan Puranjan Mishra Faculty of Civil Engineering Technology, Universiti Malaysia Pahang (UMP), Kuantan, Malaysia Nurul Nazleatul Najiha Centre for Environmental Sustainability and Water Security (IPASA), Research Institute of Sustainable Environment (RISE), School of Civil Engineering, Faculty of Engineering, Universiti Teknologi Malaysia, Skudai, Malaysia

List of Contributors

Mohd Nasrullah Faculty of Civil Engineering Technology, Universiti Malaysia Pahang (UMP), Kuantan, Malaysia Juarez Benigno Paes Department of Forest and Wood Science, Federal University of Esp´ırito Santo, Jeroˆnimo Monteiro, Brazil Kamal Kishore Pant Department of Chemical Engineering, IIT Delhi, New Delhi, India Deepak Pathania Department of Environmental Sciences, Central University of Jammu, Bagla (Rahya-Suchani), Samba, Jammu & Kashmir, India; Department of Chemistry, Sardar Vallabhbhai Patel Cluster University, Mandi, Himachal Pradesh, India Sadia Qayyum Department of Microbiology, Hazara University, Mansehra, Pakistan Yu Qin Department of Civil and Environmental Engineering, Graduate School of Engineering, Tohoku University, Sendai, Japan Ziaur Rahman Department of Microbiology, Abdul Wali Khan University Mardan, Khyber Pakhtunkhwa, Pakistan Aryama Raychaudhuri School of Infrastructure, Indian Institute of Technology Bhubaneswar, Bhubaneswar, India P.M.D. Serra Department of Electromechanical Engineering, University of Beira Interior, Covilha˜, Portugal; IT—Institute of Telecommunications, Covilha˜, Portugal Surajbhan Sevda Department of Biotechnology, National Institute of Technology Warangal, Warangal, India Swati Sharma Department of Biotechnology and Bioinformatics, Jaypee University of Information Technology, Waknaghat, India Muhammad Faisal Siddiqui Department of Microbiology, Hazara University, Mansehra, Pakistan Lakhveer Singh Department of Environmental Science, SRM University-AP, Amaravati, India T.R. Sreekrishnan Department of Biochemical Engineering and Biotechnology, Indian Institute of Technology Delhi, New Delhi, India Suryati Sulaiman Faculty of Civil Engineering Technology, Universiti Malaysia Pahang (UMP), Kuantan, Malaysia

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Shazwin Mat Taib Centre for Environmental Sustainability and Water Security (IPASA), Research Institute of Sustainable Environment (RISE), School of Civil Engineering, Faculty of Engineering, Universiti Teknologi Malaysia, Skudai, Malaysia Akash Tripathi Department of Civil Engineering, Indian Institute of Technology Hyderabad, Hyderabad, India Ihsan Ullah Department of Biological Sciences, Faculty of Science, King Abdulaziz University, Jeddah, Saudi Arabia Zahid Ullah Department of Microbiology, Hazara University, Mansehra, Pakistan Zularisam Ab Wahid Faculty of Civil Engineering Technology, Universiti Malaysia Pahang (UMP), Kuantan, Malaysia A.W. Zularism Faculty of Civil Engineering Technology, Universiti Malaysia Pahang (UMP), Kuantan, Malaysia

CHAPTER

Electrical energy produced by microbial fuel cells using wastewater to power a network of smart sensors

1

P.M.D. Serra1,2 and A. Esp´ırito-Santo1,2 1

Department of Electromechanical Engineering, University of Beira Interior, Covilha˜, Portugal 2 IT—Institute of Telecommunications, Covilha˜, Portugal

1.1 Introduction Human population is thriving, and its numbers continue to grow, even in developing countries, where it could reach 9 billion people by 2050 (WWAP, 2014). Our planet’s natural resources continue being explored and, more frequently than not, irresponsibly spent. In some developed countries, technological advancements are being thought of with environmentally friendly concerns gaining ground to disposable solutions and counterbalancing negligent behavior. Food, water, and energy supplies are also being studied for adequate management, being the most basic necessities for the sustainability and developing of life. Food cannot be produced without water; water cannot be made available without energy; and energy and electricity production are extremely limited without water. This close interconnection created a new research topic, referred to as the energy-water nexus. Considering current trends of resource depletion and climate changes, water and energy will leverage food production, resource sustainability, and technology development, balancing supply and demand (US Department of Energy, 2014). The business of energy implies more money than the water counterpart: direct costs with energy are related to exploration, treatment or generation, distribution and environmental taxes while water-related costs are mostly connected with treatment and distribution, since water is a free endogenous raw material. On top of this, energy expenditure data is richer than water expenditure and distribution (Walsh et al., 2015). By 2012, around 700 million people did not have access to an improved water source, while 1.3 billion people had no access to electricity (Walsh et al., 2015; Halstead et al., 2014). All these factors have a significant influence on policies, ensuing stricter energy management procedures than water regulations. However, a closer look at energy or water expenditure data shows their relationship, and recent data points to the need of applying similar control strategies to hinder resource and water scarcity. Water is used in almost every Delivering Low-Carbon Biofuels with Bioproduct Recovery. DOI: https://doi.org/10.1016/B978-0-12-821841-9.00002-5 © 2021 Elsevier Inc. All rights reserved.

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CHAPTER 1 Electrical energy produced by microbial fuel cells

sector of human activity, but from them all, the energy sector comes in second. Only water use for irrigated cultures comes first. In Europe, in 2015, of 247,000 million cubic meters of water, 44% was spent on energy production. For energyrelated activities, water is used for fuel extraction, refining, and processing. It also plays a significant part on the cooling of thermal power plants, accounting for 50% of the total of freshwater withdrawals in the United States and over 10% in China. It is also vital for bioenergy and biofuel feedstock crops, an energy solution easily believed to be eco-friendly and that is estimated to achieve 7.5% of global electricity production by 2050 (Walsh et al., 2015; Halstead et al., 2014; Wakeel et al., 2016; Hightower and Pierce, 2008). Water extraction, distribution, treatment, and disposal are very different in countries with and without limited energy access. By 2025, 800 million people will live in water-scarce regions and around 65% of the global population will live in severe water stress conditions. By 2050, water used for irrigation will be surpassed by the water withdrawn for energy, industrial processes, and municipal applications. Energy is needed to extract and convey water, from source to destination, and depends mostly on topography, distance and the relationship between the source’s volume of water and the amount needed at the destination (small aquifers imply more energy for water pumping). Water treatment also deals with large amounts of energy to convert wastage, rain, or seawater to a useful and safe version, both for humans and for the environment: the dirtier the water, the more energy is needed to clean it (Walsh et al., 2015; Halstead et al., 2014; Wakeel et al., 2016; do Ambiente, 2012; Copeland, 2014). The concept of energy-water nexus was first explored by Gleick (1994). For some time, since then, several events have highlighted the need for stricter regulations in water management. In 2001, there was a significant period of water shortage in California, in an event known as “The Californian Energy Crisis.” Political decisions taken at the time favored short-term human comfort in detriment for the environment and a more sustainable water management. This incident, and the likes of these, show that if no policies are applied on water management, the onset of water scarcity will happen sooner. China, the Middle East, North Africa, and Spain are the regions at most peril. Simple measures like efficient water appliances and reduced leaks in water distribution are a good way to start. Sourcing water body far from its destination, more energy will be spent on water transport. This is happening all across Europe, making water supply more energy intensive in these territories than in Asian countries. Trailing this, governments should invest on augmenting the water supply with additional sources, treating and reusing stormwater, and investigating technological solutions for freshwater production, with reverse osmosis water recycling systems and desalination plants. On power plants, alternative water sources can be used for cooling and wastewater can also be used for energy recovery (Walsh et al., 2015; Wakeel et al., 2016; Hamiche et al., 2016). In Europe, legislation is in place for energy management, with the European directive 2012/27/EU (the “Energy Efficiency Directive”) and the adoption of the

1.1 Introduction

ISO50001. This directive establishes energy utilization targets, enforces periodical energy audits and requires sustained upgrades on energy management and delivery tools. For water, similar directives are being studied and implemented. The water footprint assessment is a concept, very similar to the carbon footprint, used to quantify the water used based on life cycle assessments (LCA) and implemented with an international standard, the ISO14046. The ISO14046 evaluates the water expenditure and its impact on the environment in different life cycles stages, informing users and industry players of the water impacts of their activities and choices (Walsh et al., 2015; International Organization for Standardization, 2017). By adopting the ISO50001, several organizations showed significant decreases on energy and water usage levels as well. In Ireland, for instance, the University of Cork, which adopted the standard in 2011, saw an 18% reduction of the spent water, though an increased student activity was registered during that year. This decrease could have been higher if the standard had similar measures for water related issues, like leakages, old pipes, faulty meters and operation of sanitary facilities in low occupation periods. The implementation of this same standard by Coca Cola resulted in a 10% decrease in water and 16.5% in energy consumption (Walsh et al., 2015; Johnson et al., 2012). Following the previous background discussion, this chapter discusses the viability of producing electricity as a byproduct of wastewater processing. This discussion allows to demonstrate that the extracted energy can be made useful to power sensors that integrate a network of smart transducers compatible with the IEEE1451 standard. This is a clear example of the interconnection that is possible to achieve between two essential goods, such as energy and water: or the water-energy nexus. After the introductory section, the chapter describes the operation of a microbial fuel cell (MFC) and the energy conversion process associated with it, which uses wastewater to produce energy, while, at the same time, removes the organic content of the water. Section 1.3 demonstrates the ability to simultaneously obtain two products of high financial and environmental value: electricity and water with a low organic matter content. A wastewater treatment plant can thus be seen not only as an energy consumer, but also as having an energy production potential associated with its operation. An MFC’s operation efficiency depends on the its operation mode. As far as substrate flow is concerned, there are two possibilities: batch processing or continuous processing. For the MFC’s batch operating mode, a maximum operating point is identified in Section 1.3. The efficiency of the conversion process depends of several factors: the electrical load observed by the MFC, its temperature of operation, or organic matter availability to the bacteria are a few examples. These parameters will change continuously with time and contribute to change the operating point continuously. To increase the extraction of energy from the MFC, a tracking mechanism, that adjusts the electrical load to the MFC’s internal impedance, needs to be adopted. Proper regulation and storage of the harvested energy will support the operation of a smart sensor network.

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In this chapter, the energy production process is associated with an artificial wetland, as what occurs inside the MFC is similar to some stages in wastewater treatment processes. On the other hand, this type of infrastructure is located outside large urban centers. In these places, energy availability is reduced, which is why the production of energy in situ is an asset. At the same time, in terms of monitoring, the geographical area to be covered is vast. Considering the previous arguments, the chapter proposes, in Section 1.4, the use of the energy produced by the MFC to power wireless sensor nodes that integrate a sensor network, and that collect the information necessary for the management of the treatment process. The sensor network thus becomes energy independent as it obtains its power from the process itself. This avoids the need to use conventional batteries, and their replacement work, which can be difficult to perform, due to the high number of network nodes, and because they can be placed in difficult or dangerous locations inside the treatment plant. The application scenario will illustrate what data needs to be acquired at the wastewater treatment plant (water levels, pH, ORP, O2), and then introduce the IEEE1451 standard and its usefulness in the development of the solution.

1.2 Microbial fuel cells 1.2.1 Microbial fuel cells theoretical analysis Balance between energy production and expenditure is fragile. Although an overall energy surplus is desirable, energy deficit conditions are far more prone to occur. The development of wastewater use alternatives, namely for energy production, can tip the scales. MFCs began being explored with the knowledge gathered from traditional fuel cell technology and the previously mentioned goal. These devices operate on oxidation-reduction reactions, manipulated in a way that the produced electrons can be redirected for energy production. Fuel cell research began with the works by William Grove in 1839. Only after 1950 did the technology reach a development level adequate for industry use. Specifically, for the first American space programs, followed by the same programs in Japan and in Europe. Environmental concerns steered studies on efficiency optimization and reduced emissions in more recent years. As these areas entail different knowledge areas, a plethora of fuel cell typologies rose. The electrolyte nature ultimately determines the fuel cell type, since all fuel cells share the same functional elements: an anode, a cathode, and an electrolyte (that can be solid or liquid) (Larminie and Dicks, 2003; Scott et al., 2016). MFCs are one of those specific fuel cell types where the electrolyte can be any wastewater type: domestic, piggery wastewater, brewery wastewater, and dyeing wastewater are a few examples (Serra and Espirito-Santo, 2016). For MFCs, the anode also has a major difference from other fuel cell types: it must be colonized by exoelectrogenic anaerobic bacteria. In fact, no useful energy

1.2 Microbial fuel cells

production is possible from MFCs before the anodic biofilm is adequately developed and matured. The bacterial colony will be solely responsible for converting the wastewater organic source (biomolecules as glucose and acetate) to energy. The reactor applies conditions, anaerobiosis, such that bacteria need both the electrodes to finish the metabolic pathway from which they produce energy. The overall reaction can be described as follows (Serra et al., 2020a): 1. The electrolyte organic source is sequestered by bacteria. 2. Through anaerobic respiration, the organic source is converted to carbon dioxide and protons, with concomitant electron production. 3. The final acceptor of these electrons (O2) is made available at the cathode; the anode and cathode are connected outside of the electrolyte, and the cathode material is chosen so that it has a higher potential than the anode; this ultimately forces the exoelectrogenic bacteria to deliver the electron to the anode. The image on Fig. 1.1 presents a graphical overview of this process, highlighting the overall chemical reaction. There are several types of MFCs, sorted by reactor type or electrode disposition [Serra et al. (2020b) to be published in Instruments and Experimental Techniques, Springer]. The use of single-chamber reactors provide a more attractive solution for increased power generation since decreased internal resistances have been registered in such configurations (Fan et al., 2008). The best performance pair for the electrodes is to have a brush anode and a planar cathode. The brush shape greatly increases the reaction area, while the planar cathode is successful in guaranteeing the reactor watertightness. This set is also very useful for fixing the electrode distance, which has been found to be optimized, for energy production, for 1 cm between the last brush bristles and the cathode surface

FIGURE 1.1 A graphical representation of a microbial fuel cell with a detail of the cathode composition. AC, Activated carbon; CB, Carbon black. The governing equation of the operation process is also presented, at right. Details on this can be found in Serra et al. (2020a).

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CHAPTER 1 Electrical energy produced by microbial fuel cells

(Watson and Estadt, 2007). When relating the cathode production cost, durability and power production capabilities [as it is the bottleneck element for MFC power production (Fan et al., 2008)], the activated-carbon air cathodes produced in (Yang et al., 2014) present a satisfactory compromise. The research team of this work is currently working with single-chamber air-cathode reactors, as described in Serra et al. (2020a), that follow the aforementioned recommendations.

1.2.2 Energy extraction from microbial fuel cells 1.2.2.1 General principle The maximum voltage produced by a fuel cell corresponds to its electromotive force, or Eemf , derived from the Nernst equation: Eemf 5 Ecat 2 Ean

(1.1)

In Eq. (1.1), Ecat represents the cathode potential and Ean the anode potential. To determine these potentials, the reaction equation, the concentration of each oxidizing and reducing agent, and the specific temperature are needed. When known, the electrode potentials can be deduced from Eqs. (1.2) and (1.3): 0 Ecat 5 Ecat 2

RT lnðLÞ nF

(1.2)

0 Ean 5 Ean 2

RT lnðLÞ nF

(1.3)

and

0 0 where, Ecat and Ean represent the electrode potential in standard conditions (298.15K, 1 bar pressure, 1 M for all species), R corresponds to the universal gas constant, T is the temperature in Kelvin, and L is the reaction quotient. For MFCs, considering a pH of 7, and the general governing equation in Fig. 1.1, the theoretical Eemf is 1.1 V. This theoretical Eemf value, however, cannot be achieved because MFCs, as other fuel cells, inherently lose energy due to, for instance, unbalanced proportions between reactants and products, the reversibility of the oxidation-reduction reaction, the cell’s internal resistance—a result from the materials and geometries chosen for the reactor, the bacterial profile, and the substrate composition. All these variables play a role on the MFC’s losses and introduce ample variability and control degrees. The MFC energy losses need to be minimized to increase the device’s energy production. To do so, they first need to be identified and measured. This step is crucial if efficient adjustments are to be applied. Measures of the cell’s voltage and current with respect to external resistance and electrode’s potential as a function of time are keystone for such studies. Polarization reports are the most fundamental studies that can be conducted on an MFC. The cell can be analyzed all altogether or studies can be made independently to each electrode. For probing the cell, using a potentiometer and a voltmeter, though simple and extremely

1.2 Microbial fuel cells

accessible, produces low detailed data and can be a cumbersome process, whether due to the lack of precision in the magnitude of changes in the external load, or due to the lack of time precision of those changes. A digital potentiometer and a digital-to-analog converter (DAC) can help improve the reliability of such a method even more when combined with a microprocessor. Data retrieved through this method will allow for an adequate cell electric characterization, although not useful if the analysis of bacterial changes are of interest. When needing to conduct these studies, the use of a potentiostat is more adequate. The results of using this measuring instrument go well beyond the electrical characterization of an MFC and can also contribute to the electrochemical study of the cell’s electrodes and microbial communities. All and all, a potentiostat can retrieve the electrochemical activity of microbial strains, determine the standard redox potentials of redox active components, test the performance of cathode materials, acquire polarization curves, quantify the (overpotentials and ohmic) losses of an MFC, conduct current interrupt techniques, electrochemical impedance spectroscopy, linear, differential pulse, cyclic voltammetry, and chrono studies, namely chronoamperometry and chronopotentiometry (Logan et al., 2006; Logan, 2012; Zhao et al., 2009). Nonetheless, the potentiostat usefulness and the following discussion will consider polarization studies conducted with a potentiometer and a DAC on a mbed LPC1768 due to their price, accessibility, and ease of use. A typical polarization curve is shown in Fig. 1.2. This curve can be built by applying a single-cycle or multiple-cycle method. A single-cycle method is used when several resistance changes are made during on batch of wastewater, or if the cell is operated in continuous flow. Multiplecycle methods are applied in fed batch reactors and imply a resistance change per batch, where the number of batches will be dependent on the resolution needed for the polarization curve. A common occurrence found in polarization curves is power overshoot. A power overshoot behavior can be identified after the maximum power point in power graphs, when two different power density values can be traced back to the same current density. It is usually associated with the anode and immature biofilms; the anodic community is unable to maintain the increasing electron demand; or inadequate measurement procedures with resistance

FIGURE 1.2 An example of a polarization curve.

7

8

CHAPTER 1 Electrical energy produced by microbial fuel cells

changes too quick that do not allow the biofilm to adjust. This can happen for any reactor type, geometry, or even for reactors operated as a stack. To bypass such event, multiple-cycle methods are the best choice, since, provided the adequate readiness of the biofilm, full acclimation to the external load on the reactor is guaranteed (Logan, 2012; Prakash et al., 2010; Vicari et al., 2017; Watson and Logan, 2011; Boghani et al., 2013; Winfield et al., 2011; Hong et al., 2011). The weight and discussion of each of the MFC losses identifiable in a polarization curve can be found in (Serra et al., 2020a). Polarization curves are fundamental to adequately compare the performance of different MFC setups.

1.2.2.2 Improving energy production from microbial fuel cells The power production capabilities of MFCs are dependent on a myriad of factors, namely on reactor and electrode characteristics, and substrate composition. To make an unblemished comparison, the next three subsections and respective assumptions are supported on polarization data conducted on two types of reactors, pictured in Fig. 1.3: small volume reactors, Rx, and big volume reactors, RBigx, where x denotes the reactor number. The team has access to six small type reactors and two big volume reactors. Both reactors share the same electrodes: a carbon brush anode and an activatedcarbon air-cathode, as described in Section 2.1. Their inner chamber also has the same rectangular shape although with different sizes. The shape was kept in order to maintain the best performing electrode distance. Rx type reactors have an empty bed volume of 28 mL, while the RBigx topology accomplishes almost nine times that volume, reaching 250 mL. Another difference between the two reactor types is on the electrode number and area. The small MFCs have one electrode of each type, with a cathode area of 7 cm2. The bigger version has six interconnected anodes—disposed in a pentagon shape with a central point—and also a single cathode. The interconnected anodes work as a single electrode and the large cathode has a total area of approximately 64 cm2. This data is summarized in Table 1.1.

FIGURE 1.3 On the left (A), a picture of one of the small volume (Rx) MFCs. On the right (B), the picture of a big volume reactor (RBigx). MFC, Microbial fuel cell.

1.2 Microbial fuel cells

Table 1.1 Summary of the microbial fuel cell types used for power improvement studies. Reactor type

Anode (number)

Cathode (number)

Cathode area (cm2)

Empty bed volume (mL)

Small: Rx

Carbon brush (1) Carbon brush (6)

Activatedcarbon aircathode (1)

7

28

64

250

Big: RBigx

Electrode distance 1 cm, measured from the anode’s last bristles

In total, the number of reactors available to the team is six Rx (R1 to R6) and two RBigx (R7 and R8). The data for building the polarization curves was retrieved by the multi-cycle method, through voltage monitoring on a set of seven different resistances connected between each electrode of an individual MFC (or between electrodes of different reactors when testing for series and parallel associations). The load set is comprised of resistances of 1000, 500, 200, 100, 68, 50, and 20 Ω. The DAC on an mbed LPC1768 was programmed for data acquisition every 2 minutes. The retrieved datapoints were further processed in MATLAB 2018b. The values used for every polarization curve correspond to the maximum value retrieved from the hourly average of the 2 minutes acquired voltage values. This processing step guarantees an adequate shielding from measurement artifacts that may originate on power fluctuations of microprocessor and/or load connection defects. The trials conducted for continuous versus intermittent mode of operation are not polarization curves and use the 2 minutes acquired values. The discussion on this trial type will follow in the appropriate section. Whichever the trial, the reactors were always fed with an artificial wastewater (AW) preparation, a 50 mM sodium phosphate buffer at pH 7. The feeding solution follows the composition:

• • • • •

CH3COONa—1 g/L; NaH2PO4.2(H2O)—3.12 g/L; Na2HPO4—4.26 g/L; NH4Cl—0.31 g/L; and KCl—0.13 g/L.

A solution with such composition has around 9.8 J of energy per mL (Eq. 1.5), which can be proved by using the organic source concentration, the equation describing the energy production process (Eq. 1.4) and the individual standard Gibbs free energy of formation, displayed in Table 1.2: 1 CH3 COO2 ðaqÞ 1 2O2 ðgÞ-2HCO2 3 ðaqÞ 1 H ðaq



This value will be used to accurately pinpoint the efficiency of each trial.

(1.4)

9

10

CHAPTER 1 Electrical energy produced by microbial fuel cells

Table 1.2 Summary of the standard Gibbs free energy of formation per compound of Eq. (1.4). Compound

HCO2 3

H1

O2

CH3 COO2

ΔG0f (kJ/mol)

2586.77

0

0

2369.31

ΔG0r 5 2 3 ð 2586:77Þ 1 369:31 5 2 804:23: kJ=mol

(1.5)

Using CCH3 COONa 5 1g=LXE 5 ΔG0r 3 nH3 COONa ‘E 5 9:8J

1.2.2.2.1 The benchmark: polarization curve on a small volume microbial fuel cell The polarization trial on Fig. 1.4 has been retrieved on a small volume reactor, R6. The anodic biofilm has 674 days and the cathode applied was freshly produced, without any aerobic biofilm deposition. The inoculation procedure followed the methodology available on Serra et al. (2020a). This data will be used to benchmark all the other trials. The analysis of the data on Fig. 1.4 shows that a maximum power density of 822 mW/m2 (0.58 mW) was produced at 0.34 mA/cm2 or 100 Ω. By applying the maximum power transfer theory, this load value can be used to trace the reactor’s internal resistance (Serra et al., 2020a). A lower power density than the curve would anticipate is noticeable at about 0.4 mA/cm2. This, however, is not corroborated by the total energy extracted, as shown in Fig. 1.5. The analysis of the total extracted energy per load, considering trials run until substrate depletion, also shows that the maximum energy harvested does not occur at the load for maximum extracted power. In fact, between the two loads (200 and 100 Ω), more energy is extracted at the lowest current value. The difference corresponds to approximately 0.05 J/h, where the trial for 200 Ω run for 21 hours and the trial for 100 Ω for 17 hours, both until current values remained over 0.1 mA. The maximum efficiency of the reactor was determined to be 8.9%, with 24.5 J extracted from 28 mL of the AW (274.4 J).

1.2.2.2.2 Increasing the size of the reactor To determine if an increase in the reactor volume can be related with an increase in produced power, a polarization trial was run on a big volume reactor, RBig8, and is presented with Fig. 1.6. As with the trial on R6, the load was uninterruptedly connected between the electrodes until substrate exhaustion. In comparison with the small volume reactor, all the voltage values are higher, although current and power density figures are lower. As Table 1.3 shows, the same cannot be said for absolute power and current values, which present higher values than the small volume reactor for loads under 100 Ω, inclusive. The curve shape hints that the polarization run load values were not adequate to find the reactor’s internal resistance and, consequently, its maximum power production

1.2 Microbial fuel cells

Polarization curve for benchmarking: small reactor, R6.

0.7

900 800

0.6

Voltage (V)

600 0.4

500

0.3

400 300

0.2

Power density (mW/m2)

700 0.5

200 0.1

100

0

0

0.1

0.2

0.3

0.4

0.5

0 0.6

Current density (mA/cm)

FIGURE 1.4 Polarization curve for a small volume reactor, R6, with a fresh cathode and an anodic biofilm of 674 days (over one-and-a-half years).

Power density and total energy for the benchmarking trial

25

900

200 : 500 :

Total energy (J)

68 :

700 50 :

15

600 10 1k :

500

Power density (mW/m2)

800

100 :

20

20 : 5 400

0

0.07

0.12

0.23

0.34

0.4

0.48

0.57

300

Current density (mA/cm2)

FIGURE 1.5 Stacked view of the power curve and extracted energy values per load for the small volume reactor, R6.

capabilities. The lowest load value on the set, 20 Ω, seems to be an overestimation of the reactor’s internal resistance. Nevertheless, and for 20 Ω, the maximum power found for the trial was of 487 mW/m2 (3.1 mW). As Fig. 1.7 clearly represents, the total extracted energy was higher for 68 Ω, 207 J, which also corroborates the conclusion for the trial with the small reactor.

11

CHAPTER 1 Electrical energy produced by microbial fuel cells

Polarization curve for big volume reator: RBig8.

0.8

500 450

0.7

350 300 250

0.5

200 0.4

150

Power density (mW/m2)

400

0.6

Voltage (V)

12

100

0.3

50 0.2

0

0.02

0.04

0.06

0.08 0.1 0.12 Current density (mA/cm )

0.14

0.16

0.18

0 0.2

FIGURE 1.6 Polarization curve for a big volume reactor, RBig8, with a fresh cathode and an anodic biofilm of 415 days (over 1 year).

The volume relationship between the two reactor types is ninefold. However, neither power nor extracted energy increased in that proportion. In fact, power density decreased and the reactor performance at the maximum energy extraction was 5.8% lower. At the maximum power production load, the difference was even 3.3% higher, at 9.1%. No ninefold proportion was found between the reactor volume increase and any other energy production parameter, although the anode number and cathode surface were increased.

1.2.2.2.3 Series and parallel association Another strategy that can be used to improve the energy production levels from MFCs is the series or parallel association of two reactors. Theoretically, considering two reactors as two equal voltage sources, their series association will improve the total voltage of the set, while the parallel grouping will contribute to higher current levels. The choice of the best association will ultimately rely on the application. Nevertheless, and considering the 100 Ω internal resistance of the small reactors found with the benchmarking trial, the series association is expected to produce the highest power levels for 200 Ω and the parallel combination at 50 Ω. According to this expectation, series and parallel trials were conducted only for three load values: 200, 100, and 50 Ω. The biofilm for these trials has the same time, 674 days. The series trial datapoints are superimposed on the small MFC polarization curve in Fig. 1.8. The polarization curve is presented with absolute values of power and current since no assumption can be made on the reference area for reaction: there are no guarantees that the reactors performed the same way and, therefore, the cathode area cannot be doubled on that assumption.

Table 1.3 Summary on the data used to build the plots supporting the conclusions related with power improvement in trials with bigger volume reactors, series and parallel association and commuted load connection. Variable

Unit

Trial Small volume

Anodic biofilm time

Days

674

Cathode

Days

0

Cathode area

cm2

Voltage at max. power|max. energy

V|V

Current at max. power|max. energy

Big volume

Series association

Parallel association

415

674

7.07

63.62

7.07

0.24|0.33

0.25| 0.40

0.49|0.49

0.24|0.39

mA| mA

2.41|1.64

12.42| 5.81

2.45|2.45

4.81|1.96

Power at max. power|max. energy

mW| mW

0.58|0.54

3.10| 2.31

1.20|1.20

1.15|0.77

Max. total energy extracted|available

J|J

24.55| 274

207| 2450

39.71|548

56.19|548

Max. performance absolute|versus benchmark or continuous operation1

%

8.95|-

This table should not be used as reference without the text referring to it. a Commuted load operation. b After commuted load operation.  benchmark. 1 continuous operation.

8.45| 8.95

7.24|8.95

10.24|8.95

Interrupted load 5 min

10 min

20 min

30 min

4.99|6.26

5.04|6.42

4.21|5.54

5.54|7.83

From 0 to 4 ha 6.94|7.28 From 4 to 8 hb

From 0 to 4 ha 7.45|7.30 From 4 to 8 hb

From 0 to 4 ha 6.11|6.15 From 4 to 8 hb

From 0 to 4 ha 9.26|8.23 From 4 to 8 hb

1.82| 2.421 From 0 to 4 ha 2.53| 2.661 From 4 to 8 hb

1.84| 2.341 From 0 to 4 ha 2.75| 2.661 From 4 to 8 hb

1.54| 2.021 From 0 to 4 ha 2.23|2.24

2.02| 2.861 From 0 to 4 ha 3.38|3.00

From 4 to 8 hb

From 4 to 8 hb

CHAPTER 1 Electrical energy produced by microbial fuel cells

Power density and total energy for the big volume reactor.

250

500 450

68 :

100 :

50 :

400 350

20 :

200 :

150

300 250 100

500 :

200

1k :

50

Power density (mW/m2)

Total energy (J)

200

150 100

0

50 0.01

0.02

0.04

0.07

0.09

0.11

0.2

Current density (mA/cm2)

FIGURE 1.7 Stacked view of the power curve and extracted energy values per load for the small volume reactor, RBig8.

1.4

0.6

1.2

0.5

1

0.4

0.8

0.3

0.6

0.2

0.4

0.1

0.2

0

0

0.5

1

1.5

2

2.5

3

3.5

4

Power (mW)

Polarization data comparison between a single small reactor and a series association of 2 small volume reactors.

0.7

Voltage (V)

14

0 4.5

Current (mA)

FIGURE 1.8 Polarization curve of the small reactor (R6) and datapoints for series association of two small volume reactors (R5 and R6).

Overall, the series association of two reactors produced higher voltage values, except for 50 Ω. As far as current is concerned, the values were approximately the same for a single reactor or two series connected MFCs. These datapoints seem to point that, for lower loads, the series behavior approaches the performance of a single MFC. As far as power is concerned, due to the higher voltages resulting from the series association, values are also higher for higher loads. In fact, for the 200 Ω load—corresponding to the series association of two reactors with an internal load

1.2 Microbial fuel cells

10%

Performance and total energy comparison between a single small reactor and a parallel association of 2 small volume reactors.

1.4

9% 1.2 8%

Perfomance (%)

6%

0.8

5% 0.6

4% 3%

Power (mW)

1

7%

0.4

2% 0.2 1% 0%

0 50

100

200

Load (:

FIGURE 1.9 Comparison of the power and extracted energy values, per load, for a single reactor and a series association.

of 100 Ω—the value is 52% higher than the one achieved with a single reactor (1.2 mW vs 0.58 mW). For two reactor series grouping, power seems to double when compared with power harvested with a single reactor. When analyzing the extracted energy, the available energy considered for calculation was two times that of a single reactor. This has produced the plots on Fig. 1.9. Irrespective of the load value, the performance of this trial was always lower than the performance of a single reactor. The biggest performance difference was 73.7%, observed at 100 Ω. The maximum energy harvested was 39.7 J of a combined total of 548.8 J, for the same load as the maximum power (200 Ω). On the parallel association of reactors, a power increase is also predictable due to increased currents. The same consideration for power and current presented for building the series curve is applied to Fig. 1.10. In this mode of operation, both voltage and current figures have increased. The best performance is found at 50 Ω, substantiating the theoretical prediction. All datapoints present significative gains when compared to the single reactor, with absolute power improvements ranging from 29.75% at 200 Ω to 50.09% at 50 Ω. Data on Table 1.3 provides precise values that authenticate the above data and plot. As with series grouping, also parallel merging of reactors leads to increased power, with simultaneous increases in voltage and current. A similar behavior can be found for the performance of the reactor, as showed by the plot on Fig. 1.11. The best performance of the trial was found at the load for maximum power, 50 Ω, and corresponds to a 19.7% improvement from the single reactor. At 200 Ω the improvement was smaller, around 12.6% while at 100 Ω the single reactor outperformed the parallel association of MFCs.

15

CHAPTER 1 Electrical energy produced by microbial fuel cells

Polarization data comparison between a single small reactor and a parallel association of 2 small volume reactors.

1.4

0.6

1.2

0.5

1

0.4

0.8

0.3

0.6

0.2

0.4

0.1

0.2

0

0

0.5

1

1.5

2

2.5

3

3.5

4

4.5

5

Power (mW)

Voltage (V)

0.7

0

Current (mA)

FIGURE 1.10 Polarization curve of the small reactor (R6) and datapoints for parallel association of two small volume reactors (R5 and R6).

11%

Performance and total energy comparison between a single small reactor and a parallel association of 2 small volume reactors.

1.4

10% 1.2 9% 8%

1

7% 0.8

6% 5%

0.6

Power (mW)

Perfomance (%)

16

4% 0.4

3% 2%

0.2 1% 0%

50

100

200

0

Load (:)

FIGURE 1.11 Comparison of the power and extracted energy values, per load, for a single reactor and a parallel association.

A power comparison between single reactor, series and parallel association is presented in Fig. 1.12. This complete representation highlights the clear dominance of the 50 Ω parallel trial, with increased power and current when compared to every other trial. It is also very noticeable that lower loads improve the produced current. When comparing the performance of each trial, using Fig. 1.13, it becomes clear that the 200 Ω load leads to greater energy extractions from the reactors, particularly

1.2 Microbial fuel cells

Polarization data comparison for a single small reactor, a series and a parallel association of 2 small volume reactors.

1.2

Power (mW)

1

0.8

0.6

0.4

0.2

0

0

0.5

1

1.5

2

2.5 Current (mA)

3

3.5

4

4.5

5

FIGURE 1.12 Comparison between the power production capabilities of a single reactor, a series, and a parallel association.

11

Performance comparison for a single small reactor, a series and a parallel association of 2 small volume reactors.

10 9

Performance (%)

8 7 6 5 4 3 2 1 0

50

100

200

Load (:)

FIGURE 1.13 Comparison of the power and extracted energy values, per load, for a single reactor and a parallel association.

for parallel associations. For 100 Ω, a single reactor has a better energy extraction than any other association and the parallel grouping is highlighted again at 50 Ω. Considering both the power and performance studies, for small reactors, the best way to use two MFC reactors for energy production is to have them connected in parallel and subject to a load which value equals the parallel equivalent of their

17

18

CHAPTER 1 Electrical energy produced by microbial fuel cells

internal resistances. This work should be further investigated for associations of more than two reactors. Also, other studies have found that these associations are subjected to voltage reversal phenomenon, which limits their application capabilities (Oh and Logan, 2007; Boghani et al., 2014). Other studies have determined that an advantageous way to explore these associations is to have them done intermittently, rather than continuously (Khaled et al., 2016; Nguyen et al., 2019; Boghani et al., 2017).

1.2.2.2.4 Continuous versus intermittent mode of operation The last researched strategy for power optimization of MFC reactors was identified when open circuit voltage conditions were imposed, for the same AW batch, before load connection, as reported in Fig. 1.10 of Serra et al. (2020a). The plot shape suggested a capacitive behavior of the reactor, more evident when the external load applied was close to the cell’s internal resistance value, and even more when if applied in successive trials. The work on this subject is herein presented. Trials were conducted for 50% duty cycles and ON times of 5, 10, 20, and 30 minutes. The trial consisted in feeding the reactor with fresh AW and connect a 100 Ω load for 30 minutes before starting the interrupted connection investigation. After 30 minutes, the load was connected and disconnected from the MFC with different ON times per wastewater batch. For instance, a ON time of 5 minutes was tested for a wastewater batch. A new ON time of 10 minutes was conducted only when the reactor’s volume was completely replaced by a new AW batch. The period for comparison was 4 hours and datapoints were acquired at 2 minutes intervals. The energy calculus was the cumulative sum of the 2 minutes differential energy for 30 minutes. When evaluating the interrupted load trials, the comparison was made for the same operating period than an uninterrupted load connection trial. The plots on Fig. 1.14 start the voltage comparison discussion. The interrupted load connection seems to induce an overvoltage. This phenomenon may be used simultaneously with parallel association of reactors and may lead to further improvements in power production. To adequately measure the impact of this overvoltages on energy production, a study of the energy per trial, as presented in Fig. 1.15, is needed. An increase in the ON times leads to a decreased performance of the reactor. This trend is not yet clear for ON times inferior to 5 minutes, where research is needed. This plot also clarifies that, although there are significative overvoltages—and higher for higher ON times—the extracted energy is always inferior to the energy extracted in continuous operation. The impact of this operation mode on continuous operation was also investigated. After 4 hours of the trial beginning, the uninterrupted operation mode was applied. The data groups of ON time periods were maintained for reference. Results are plotted in Fig. 1.16. The comparison was made for the same operation time, from 4 to 8 hours. For 4 hours after the load commutation, only the 5 minutes ON time trial produced

1.2 Microbial fuel cells

5 min commutation

0.3 0.2 0.1 0 00:00:00

01:00:00

02:00:00

10 min commutation

0.4

Voltage (V)

Voltage (V)

0.4

03:00:00

0.3 0.2 0.1 0 00:00:00

04:00:00

01:00:00

Time (h)

Voltage (V)

Voltage (V)

0.2 0.1

01:00:00

02:00:00

04:00:00

30 min commutation

0.4

0.3

0 00:00:00

03:00:00

Time (h)

20 min commutation

0.4

02:00:00

03:00:00

04:00:00

0.3 0.2 0.1 0 00:00:00

Time (h)

01:00:00

02:00:00

03:00:00

04:00:00

Time (h)

FIGURE 1.14 Voltage datapoints acquired at 2-min intervals for 4 h for R6 in continuous and interrupted load operation. The interrupted load operation was tested for ON times of 5 (top left), 10 (top right), 20 (bottom left), and 30 (bottom right) min, always with a 50% duty cycle.

8

Performance and total energy comparison between commuted and uninterrupted load operation.

7

0%

–5%

6

5 –15% 4 –20%

Perfomance (%)

Total energy for 4 h (J)

–10%

3 –25% 2

–30%

1

0

–35% 5

10

20

30

Commutation time (min)

FIGURE 1.15 Comparison of the total energy and performance, per ON time and for 4 h of operation, for commuted and uninterrupted load operation of a single reactor.

19

CHAPTER 1 Electrical energy produced by microbial fuel cells

10

Performance and total energy comparison for the operation after commuted and uninterrupted load operation. 12%

9 10% 8 8% 7 6% 6 4% 5 2%

4

Perfomance (%)

Total energy for 4 h (J)

20

0%

3

2

–2%

1

–4%

–6%

0 5

20

10

30

Commutation time (min)

FIGURE 1.16 Comparison of the total energy and performance, per ON time for 4 h after the commuted operation (from 4 to 8 h).

lower energy levels. All other ON times lead to equal or increased performances with the 30 minutes ON time leading the improvement with 12.57%. The research work with interrupted load connection has allowed to determine that the reactor performance is not enhanced for ON times over 5 minutes if continuously operated in such a fashion. However, further research can be conducted to determine if this operation mode allows to operate an MFC for longer periods of time with the same AW, which can ultimately be useful in wastewater treatment plants (WWTP) with conditioned access to organically enriched wastewater. An investigation path is still open for interrupted connection evaluation of ON times inferior to 5 minutes and different duty cycles, although duty cycles higher than 50% lead the reactor to continuous operation performances.

1.2.2.3 Closing remarks There are several optimization strategies for MFC energy harvesting, including, for instance, power management systems or improvements in materials, design or biological manipulation of the colony. However, the trials here explored are examples of power optimization strategies exclusively dependent on the reactor operation mode. A summary on the findings of this trials is presented in Table 1.3. Important advances in MFC technology have recently been put forward. Nonetheless, this technology is not yet ready to be implemented in WWTP as a power source for the energy intensive processes, like anaerobic digesters or

1.3 Energy production, regulation and storage

stirrers. At its current technological point, it can be of great use to power chemical and physical sensors that aid in the flow management of WWTP, providing an autonomous, reliable, and ecological solution. The research on MFC applications to real wastewater substrates or even on adapted WWTP has been increasing, giving a clear indication of the technology readiness for further investment and research. In 2019, Su et al. (2019) and Santoro et al. (2019) addressed the performance of MFCs with real wastewater: while Su et al. used piggery wastewater, Santoro et al. researched a particular type of stratified MFC to use urine substrate. As the WWTP incorporation is concerned, the most recent work was by Xu et al. (2018), continuing the work initiated by Srivastava et al. (2017) in 2017, that converted the MFC to a benthic MFC by burring the anode and leaving the cathode afloat. The team managed to extract 1.3 W/m3 in 2-L reactors. The work here presented is of importance when combined with other optimization strategies and provides a general overview of the most classical ways to improve the energy harvesting of MFCs.

1.3 Energy production, regulation and storage Before using the energy harvested from an MFC to power a smart sensor, two main problems need to be addressed. First, the voltage value is low and unregulated. Second, power output is not high enough to power a smart sensor in the operating mode continuously. The first issue is solved using a two-stage solution that, at first, raises the voltage level, to allow a DC/DC converter start operating. The second problem is solved by adopting low-power operating modes based in low-power deep sleeping mode that optimize the use of the available energy. These two problems are discussed next.

1.3.1 Energy regulation and storage The magnitude of the energy produced by an MFC is very low. It typically reaches 0.8 V when operating in open circuit, a condition where the internal impedance does not affect the cell’s voltage output because no current is flowing between the electrodes. As described in the previous section, an increase in the cell’s current leads to a drop in the corresponding output voltage, due to an increase in its internal impedance. This characteristic behavior usually makes the cell incapable of starting a DC/DC converter, which is needed to guarantee an adequate voltage regulation to adequate the cell to power a smart sensor. An answer to this issue is to use a second stage implementing a voltage raising mechanism capable of providing the necessary conditions for the start of the DC/DC converter. This second stage also helps to operate the converter in the most efficient way. The proposed solution is based on the use of a two-stage converter, as previously described.

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Since harvesting energy from an MFC, with this converter, is not a direct process, an operation control is required. The operation control is imposed by four switches: S0, S1, S2, and S3. Switches S0 and S1 are normally closed, while S2 and S3 are normally open. The commutation of these switches leads to two different operation conditions: a starting stage and a regular operation stage.

1.3.1.1 Starting stage Initially, the smart sensor cannot run because there is no power in the system. As such, all the switches are set to their default behavior. This forces the MFC’s voltage to capacitor 1 (C1), which, when reaching the blocking oscillator’s (BO) starting voltage (BOSV), charges capacitor 2 (C2). The voltage supervisor 2 (SV2) monitors C2’s voltage: when it reaches SV2’s high threshold value (SV2_H), the S2 switch closes. From that moment onwards, the DC/DC converter starts the voltage regulation from capacitor 3 (C3). The S2 switch remains closed as long as C2’s voltage remains higher than the SV2’s low threshold voltage (SV2_L). When the C3’s voltage reaches the SV3_H, SV3 commutes the S3 switch, closing it. This cascade of events finally leads to the microcontroller unit (MCU) activation. This ends the starting stage and initiates the regular operation.

1.3.1.2 Regular operation stage At this operation condition, the MCU controls the S0 and S1 switches, periodically commuting them. The capacitor C1 is the first to charge, keeping the S0 and S1 switches open. After C1 is charged, S0 is opened an S1 closed, starting C2 through the BO oscillator with the energy from C1. The MCU is master of this process, using the junction gate field-effect transistor (JFET) driver controller lines and setting the commuting values of C1’s voltage. The implemented control strategy is based on monitoring the voltage at C1. The S0 and S1 switches are managed such as the C1’s voltage varies between VC1_H and VC1_L. These boundary values are detected by the voltage comparator (VC), which compares them with the reference voltage (VREF) output from filtering the pulsewidth modulation (PWM) (PWM_REF), signal generated by the MCU. In the charging period, with S0 closed and S1 open, the VC sees VREF_H. When the higher value is achieved, the control line C1_Status triggers a low to high transition. The MCU replies to this event by opening S0 and closing S1, starting the discharge period. The reference voltage is set to VC1_L, a value resulting from filtering a MCU’s PWM. When the capacitor reaches the discharge value, the control line C1_Status triggers a high to low transition, generating a new interruption at the MCU that opens S1 and closes S0 to allow a new charging cycle for the C1 capacitor. As long as the energy stored at C2 is enough to guarantee that the DC/DC converter is able to maintain the C3 capacitor with a voltage higher than SV3_L, the MCU continues to operate and control the S1 and S0 switches. If the C2 voltage drops below SV2_L, or if the C3 voltage gets lower than SV3_L, the MCU will enter in sleep mode, after being notified by the SV2_Status and SV3_Status lines.

1.3 Energy production, regulation and storage

FIGURE 1.17 Energy module structure diagram.

A general overview of the energy module is presented in Fig. 1.17. Each of its blocks is described, in more detail, in the following subsections.

1.3.1.3 Oscillator The BO, pictured in Fig. 1.18, is composed by a transformer, a resistance, and a transistor working as an amplifier. The assembly places the transformer in a positive feedback. When the transistor is saturated, the second winding current will produce a voltage leading to a decrease in the transistor’s base current, since the magnetic flow of the secondary winding of the transformer has an inverse signal of the primary winding. In this condition, the transistor immediately enters in cutoff, rapidly increasing the voltage at the transformer’s secondary winding. The energy stored in the magnetic circuit of the transformer is transferred, through the diode, to the capacitor. When the voltage at the secondary winding of the transformer drops, the current in the primary circuit will increase again quickly, placing the transistor again in its saturation zone, conducting.

1.3.1.4 Voltage comparator The MFC operation efficiency is optimized by adjusting the charge and discharge times of the C1 capacitor. The C1’s voltage is monitored by a comparator,

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FIGURE 1.18 The blocking oscillator.

FIGURE 1.19 The voltage comparator.

displayed in Fig. 1.19, that evaluates its value against the reference voltage VREF supplied by the MCU. At any particular time when the capacitor voltage is higher than the reference value, the comparator output is set to high. At transitions, the MCU answers by commuting the S0 and S1 positions, according to the regulation strategy defined. The reference voltage level is set by a DAC, implemented by filtering a PWM signal, with 256 different levels, resulting in a 12-mV resolution. A low-pass filter is used to determine the PWM average value.

1.3.1.5 Field effect transistor driver A JFET is capable of conducting current between drain and source when the terminal gate has no voltage. To turn off an N-channel JFET a high enough negative voltage has to be applied at the terminal gate. Due to this behavior, the JFET can be designated has a normally closed switch. A charge pump uses a PWM generated by the MCU to produce the negative voltage needed to place the JFET at cutoff, opening it. Fig. 1.20 represents the field effect transistor driver and its desired operation.

1.4 Smart sensor structure and operation

FIGURE 1.20 The field effect transistor driver.

1.3.1.6 Voltage supervisor The charge status of capacitors C2 and C3 is monitored by voltage supervisors. Whenever the voltage value is higher than a preset value, the S2 and S3 switches status is updated, according to the variation direction. The LTC2935 device is responsible for this task, allowing to program the levels of commutation of the status line. The supervising circuit monitors the input voltage and sets the reset (RST) line to low when the capacitor voltage drops below the preset threshold value. When the capacitor voltage increases, and surpasses the programmed high threshold voltage in 5%, an internal timer delays the RST line reset to high in 200 ms. Several different commutation levels are configurable by using three configuration lines. While the capacitor charges, and with the supervisor configured to detect the higher voltage level (S2 5 \RST (Low); S1 5 Low; S0 5 High), the RST line is low, which sets the limit value to 3.15 V. When the higher value is reached, the RST line is triggered to high and closes the switch S1. The voltage supervisor is now configured to detect the lower limit (S2 5 \RST (High); S1 5 Low; S0 5 High), corresponding to 2.40 V. The DC/DC converter now begins its operation and the capacitor voltage decreases. When it reaches 2.40 V, the RST line is again triggered, this time to low. This forces a switch opening and the capacitor charge.

1.4 Smart sensor structure and operation The function of a smart transducer is to acquire information from the process where it is deployed, if it has a sensor element, or to actuate in that same process, if it has an actuator. As the transducer definition suggests, this element can be used to either designate a sensor or an actuator. A smart transducer has a set of characteristics that allows it to perform functions such as amplification, filtration, and data processing; communication with other elements through a network, autonomous operation, or in collaboration with other elements interconnected through the same network.

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FIGURE 1.21 General overview of the smart sensor on discussion.

The power module described in the previous section is an essential building component of a smart sensor. An MCU is the central element around which all other modules are developed. In addition to having the function of managing the power module, as already described, the MCU manages the remaining modules as shown in Fig. 1.21. The measurement chain links the sensor element to the MCU. It is through this channel that the acquired information is first amplified and filtered in the analog domain to, after being converted to the digital domain, be digitally processed by the MCU. The radio module connects the smart transducer to the network. It is through this channel that information requests are received and sent. The energy available in the smart transducer needs to be managed, taking into account aspects such as those discussed next. The data to be acquired can be obtained (1) based on a periodic sampling process over time, (2) at the request of another process, or (3) due to the occurrence of associated events. 1. In the periodic sampling process, information is acquired with a predefined sampling frequency. The measurement process is triggered periodically. It is possible to estimate the energy cost of this process. Thus, the required energy, and given that we are using an energy collection mechanism, should never be greater than what can be acquired between samples. 2. The smart transducer can receive a request for information over the network. In this situation, it is necessary to ensure that there is an adequate energy stock so that the measurement process can be flawlessly carried out. Ensuring that the request occurs in periods of time longer than what is necessary for the energy level to recover is mandatory. 3. The report of an event occurrence is the sole responsibility of the smart transducer. Therefore, an adequate energy stock must be made available for the communication to happen. Ensuring that the request occurs in periods of time longer than what is necessary for the energy level to recover is mandatory.

Acknowledgments

An example of applying the solution presented in this chapter is the case of a wastewater treatment plant. At these infrastructures, data registering is necessary: water level, temperature, pH, ORP, and O2 are some examples of chemical and physical parameters that need monitoring in order to properly meet the environmental discharge standards. In the work presented in Esp´ırito-Santo et al. (2017, 2019), a network of intelligent actuators that collects the water level in a macrophyte lagoon is described by the authors. Smart sensors are powered by MFC. Network interconnection is based on the IEEE1451 standard. This standard defines an NCAP element that acts as a gateway between the user’s network and a set of TIMs that implement smart and wirelessly connected sensors. One aspect of this rule is that each TIM is described by a set of TEDS that have all the necessary information so that an application having access to them can recognize TIM.

1.5 Conclusions The ecological awareness of today’s society is growing. The impact that the human species had on the planet in the last century has contributed to this. The result is visible in the continued degradation of natural environments, some beyond possible recovery. The study of the water-energy nexus has contributed to identify processes where there is a direct relationship between these two essential resources for humans. This chapter demonstrates that it is possible to produce energy and treat wastewater at the same time. The bacteria, in addition to degrading organic matter, if the process takes place in an anaerobic environment, produce an electric current that can be used to power an electronic device. The production of electrical energy using MFCs, despite being a simple procedure, requires a detailed study of several aspects that can range from the geometric configuration of the reactor, to the mechanisms through which it is possible to efficiently extract energy. In order for the energy produced by the MFC to be used, it is necessary to operate the cell at an optimal point. This operation can become more efficient if the management algorithm has the capacity to adapt, in time, to different operating situations. With an energy source, perfectly integrated in the process, it is possible to power a smart transducer that collects information from the process and sends it through a network, thus contributing to the optimization of the wastewater treatment process.

Acknowledgments This work was funded by the Portuguese Foundation for Science and Technology (FCT) (Grant ERANETMED/0004/2014), through the ERANETMED initiative of Member States, Associated Countries and Mediterranean Partner Countries (Project ID eranetmed_nexus14044).

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CHAPTER

Application of bioelectrochemical systems in wastewater treatment and hydrogen production

2

Santhana Krishnan1, Abudukeremu Kadier2, Mohd Fadhil Bin MD Din1, Mohd Nasrullah3, Nurul Nazleatul Najiha1, Shazwin Mat Taib1, Zularisam Ab Wahid3, Yu You Li4, Yu Qin4, Kamal Kishore Pant5, Shreesivadasan Chelliapan6, Hesam Kamyab6, Imran Ahmad6 and Lakhveer Singh7 1

Centre for Environmental Sustainability and Water Security (IPASA), Research Institute of Sustainable Environment (RISE), School of Civil Engineering, Faculty of Engineering, Universiti Teknologi Malaysia, Skudai, Malaysia 2 Laboratory of Environmental Science and Technology, The Xinjiang Technical Institute of Physics and Chemistry, Key Laboratory of Functional Materials and Devices for Special Environments, Chinese Academy of Sciences, Urumqi, China 3 Faculty of Civil Engineering Technology, Universiti Malaysia Pahang (UMP), Kuantan, Malaysia 4 Department of Civil and Environmental Engineering, Graduate School of Engineering, Tohoku University, Sendai, Japan 5 Department of Chemical Engineering, IIT Delhi, New Delhi, India 6 Department of Engineering, Razak Faculty of Technology and Informatics, Universiti Teknologi Malaysia, Jalan Sultan Yahya Petra, Kuala Lumpur, Malaysia 7 Department of Environmental Science, SRM University-AP, Amaravati, India

2.1 Introduction The world economy is powered by fossil fuels (FFs), including oil, coal, and gas, but the FFs are unsustainable and gradually depleting. In addition, burning of FFs release greenhouse gas (GHG), mainly carbon dioxide (CO2), which in turn cause global warming and acidification of the ocean (O’Connor et al., 2020). Hence, cleaner and more renewable sources of energy are being sought to power our world in a sustainable manner. Recently, hydrogen (H2) has gained tremendous potential as a fuel and energy source of the future. H2 is a green fuel with a high energy content of 142 MJ/kg and does release only water during combustion. Moreover, H2 can be produced from a variety of renewable feedstocks, mainly organic wastewater (Din et al., 2020a). Currently, 80% of H2 around the world is produced through a thermochemical process such as steam reforming of natural gas, coal gasification, and electrolysis, which mainly involve using FFs as the Delivering Low-Carbon Biofuels with Bioproduct Recovery. DOI: https://doi.org/10.1016/B978-0-12-821841-9.00003-7 © 2021 Elsevier Inc. All rights reserved.

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feedstock. This energy intensive process requires high electricity and further leads to global warming increase and generate a carbon footprint in the ecosystems (Vasylieva et al., 2019). H2 production using biomass seems to be a promising alternative technology to reduce environmental pollution and is now given high importance. Biohydrogen is established as renewable and green methods, which typically include biophotolysis, dark fermentation (DF), and photo-fermentation. Among these methods, DF is considered to be more promising because the process does not require a light source, utilizes low energy demand and various organic substrates such as carbon sources, and needs mild operating conditions (Krishnan et al., 2020). However, DF leaves behind some organic volatile fatty acid (VFA) end products such as butyric acid, acetic acid, propionic acid, and ethanol that has its own thermodynamics limitations which may result in lower H2 yield (Mishra et al., 2019). Microbial electrochemical systems (MEC) is a promising technology to produce clean and sustainable hydrogen energy and wastewater treatment simultaneously. MEC has a potential to convert organic wastewater into hydrogen and value-added specialty chemicals such as methane (CH4), ethanol, and hydrogen peroxide (Krishnan et al., 2019). Compared with other conventional methods, MEC offers high H2 yield with a small energy input of 0.4
0.5 V. Such energy requirements are lower than the energy needed for water electrolysis (50 V). In addition, it overcomes the DF limitations by converting the end products further into hydrogen (Jia et al., 2020).

2.2 MEC fundamentals and working principles The principal components of the MEC consist of anode and cathode electrodes, semipermeable membrane, electrochemically active microbes (EAB), and the power supply unit. Schematic diagram of typical two-chambered MEC and its operation is given as Fig. 2.1. In MEC, EAB attached to the anode layer degrades organic matter into protons (H1), electrons (e2), and CO2. As a part of metabolism, EAB transfers e2 into the anode surface, while H1 ions are diffused into MEC electrolyte solution (Das et al., 2020). The diffused H1 ions combine with the e2 that are traveled through a wire of cathode region to produce H2. However, this reduction mechanism is not spontaneous and it needs 20.414 V of electrical input under standard temperature, pressure, and pH (pH 5 7, T 5 25 C, PH2 5 1 atm) (Heidrich et al., 2014). A simple MEC principle is shown using acetate as an example. CH3 COO2 1 4H2 O2 HCo2 1 H1 1 4H2 2 0:414V

(2.1)

2.2 MEC fundamentals and working principles

FIGURE 2.1 Schematic diagram of typical two-chambered MEC and its operation. EAB, Electrochemically active microbes.

The Nernst equation is used to calculate the reduction potential of the half-cell reaction. Cathode reduction potential is calculated according to following equation: 0 Ecat 5Ecat 2

RT PH2 8:314 3 298 1 502 5 2 0:414V 1n 1n 2F 2 3 96; 485 ½1027 8 8 ½H1 

0 Ecat is standard electrode potential for hydrogen (0 V), R (8.314 J/K/mol) is the universal gas constant, T (K) is the absolute temperature, and F (96,485 C/mol e2) is Faraday’s constant. Anode reduction potential is calculated according to following equation: 0 Ean 5Ean 2

RT ½CH3 COO2  8:314 3 298:15 0:0169 1n 1n 5 0:187 2 5 2 0:414 5 8F ½HCO32 8 ½H1 9 8 3 96; 485 ½0:0052 ½1027 8

0 Ean 0:187 is standard electrode potential for acetate oxidation.

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The cell voltage required for MEC to generate H2 at cathode junction under these conditions are Ecell 5 Ecat2Ean 5 ð 20:4:14VÞ 2 ð 20:3VÞ 5 2 0:114V

The Gibbs free energy (ΔG) must be negative for any reactions to occur spontaneously however, the transformation of organic matter to H2 gives a positive Gibbs free energy. Hence, input voltage of . 0:114V has to be supplied in order to favor the H2 in MEC (Rousseau et al., 2020).

2.3 Electron transfer mechanism In MEC, the electron transfer mechanism (ETM) is achieved by the direct and indirect ETMs. The schematic diagram of extracellular electron transport during MEC operation is given in Fig. 2.2. During the MEC operation, the transfer of e2 from the organic substrate to the electrode is essential, and it is performed by the EAB, also known as exoelectrogens/electrogens. Such electrogens mediate electron transfer from organic substrate extracellular e2 acceptor without mediators (Mohamed et al., 2020). Diverse groups of EAB include,

FIGURE 2.2 Schematic diagram of extracellular electron transport during MEC operation.

2.4 MEC technology in hydrogen production using wastewater

Proteobacteria (Ochrobactrum, Rhodopseudomonas, Geobacter, Desulfobulbus, Arcobacter, Klebsiella, Rhodoferax), Firmicutes (Thermincola, Clostridium), and Acidobacteria (Geothrix). Direct transfer mechanism involves direct electron transfer from EAB (Geobacter, Shewanella) to the electrode (electron acceptor) through the outer membrane with the help of cytochrome and heme proteins (Aboelela et al.). The direct method also involves the transfer of e2 through soluble e2 through protrusions from cells pili, and also via diffusive shuttle compounds known as phenazines, quinines, melanin, and Flavin’s. Indirect ETM, also known as mediated ETM, employs soluble redox species as interconnecting bridges and involves no direct contact between the electrode and EAB (Flores-Estrella et al., 2020).

2.4 MEC technology in hydrogen production using wastewater Due to advancements in reactor design and operational methods, MECs are widely used in product synthesis, resource recovery, waste treatment, sensing, and electrochemical research. The waste-to-wealth concept holds immense potential in waste reduction and renewable energy production. In MEC, H2 can be produced using various organic substrates such as glucose, lactose, cellulose, acetic acid, butyric acid, and lactic acid; different types of wastewater (agro waste, domestic, industrial waste, etc.) substrates are found to be the most important factors influencing H2 production (Din et al., 2020b). DF is a complex biochemical reaction in which the organic matter is disintegrated into soluble particles in a series of steps by anaerobic fermentative bacteria. Firstly, the organics (mainly proteins, carbohydrates, and lipids) are hydrolyzed into simple monomers such as amino acids, monosaccharides, and fatty acids. Secondly, the monomers are further used by the bacteria to produce soluble intermediary organic acids (VFAs), alcohol, and CO2 by a process called acidogenesis. Finally, the organic acids are converted into H2 and acid intermediates through the process called acetogenesis. The main drawback of the DF is low conversion efficiency and high organic chemical oxygen demand (COD) levels in the fermentation effluent. Hence, the integration of MEC with DF is a great solution to overcome the limitations associated with the DF, and also, this technology enhances the high organic removal rate and energy production. Recently, many studies have proved that MEC and DF can be integrated to enhance the efficiency of simultaneous H2 production and wastewater treatment (Huang et al., 2020). Integration of DF-MEC results in the theoretical production of 12 moles H2 per mole of hexose while the DF process alone is limited to 4 moles H2 per mole of hexose (Jia et al., 2020). Acetate is the most widely used carbon source in MEC. Jeremiasse et al. (2011) used acetate in MEC with the 1 V power supply. In addition to this, various fermentable and nonfermentable substrate such as glucose, sucrose, cellulose,

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CHAPTER 2 Application of bioelectrochemical systems

butyrate, propionate, lactate, and valerate were used in MEC with the current density ranged from 1.1 to 6.2 A/m2. The maximum H2 yield of 0.46 LH2/L/d was obtained in a single-chambered alkaline MEC using glycerol as the substrate by Badia-Fabregat et al. (2019). Table 2.1 depicts the various wastewater used in MEC for hydrogen production and treatment.

2.5 Agro wastewater Agro biomass or lignocellulose biomass is the most abundant resource in the earth, and it is composed of cellulose, hemicelluloses, and lignin. Lignocellulose biomass comprises energy crops, waste biomass, and virgin biomass (grass, trees, and bushes). Lignocellulose compounds are hard to degrade by microorganisms in MEC, and hence, pretreatment is the best way to convert recalcitrant into monomers (Gil-Carrera et al., 2013). The renewable characteristics of lignocellulose biomass render agro residues as the suitable carbon source for the microorganisms in the MEC for hydrogen evolution. Combining DF with electrohydrogenesis, which is known as two-phase DF, allows greater hydrogen production rates and yields using recalcitrant lignocellulose biomass. This two-stage process includes optimal conversion of pretreated lignocellulose biomass into hydrogen, carbon dioxide, and various VFAs like acetic, butyric, propionic, formic, succinic acids, and ethanol, which is followed by the VFAs and alcohol into hydrogen (Catal, 2016). Lalaurette et al. reported 1.019 and 0.97 m3 H2/m3 d of hydrogen production rate using the effluent of lignocellulose and cellobiose substrate through two-state fermentation. Lalaurette et al. (2009) investigated hydrogen production using MEC from various sugar substrates like glucose, maltose, mannose, arabinose, and cellobiose, and he achieved the hydrogen production rate of 0.01 m3 H2/m3 d
0.10 m3 H2/m3 d (Lalaurette et al., 2009). On the contrary, Lewis et al. explained a new concept of integrated pyrolysis and MEC using switch grass as an anodic chamber substrate. The hydrogen yield of pyrolysis ranged with an average of 70% while the hydrogen yield from MEC was found to be 90%, respectively (Lewis et al., 2015).

2.6 Domestic waste water In earlier studies, MEC was examined with the acetate as the carbon source for H2 production. It is fair to assume that domestic wastewater may be used to produce hydrogen. Escapa et al. investigated hydrogen production using domestic wastewater as the substrate in the MEC and evaluated system performance in terms of the organic loading rate (OLR) and applied voltages (Vapp). Results concluded that MEC was able to reduce up to 76% of the COD of domestic wastewater. The hydrogen production rate peaked at 0.3 L/Lr/d (hydraulic retention time:

Table 2.1 Various substrates used in MEC technology for wastewater treatment and H2 generation. Wastewater

Initial COD

MEC

Electrodes

Voltage (V)

Hydrogen production (m3/m3/d)

COD removal rate

References

Potato wastewater Ethanol Effluent Industrial wastewater Domestic wastewater

2.6 g/L

Single chamber fed batch Single chamber fed batch Single chamber fed batch Multiple chamber MEC cell cassettes, continuous Multiple chamber MEC cell cassettes, continuous

Ammonia graphite brush/carbon/pt. Graphite brush/ carbon/pt. Heated graphite brush/MoS2 Stainless steel wool

0.9

0.74

76%
87%

Kiely et al. (2011)

0.8

2.1

70%

Lu et al. (2009)

0.7

0.18

2.8 kg-COD/m3 d

Tenca et al. (2013)

0.9

0.17

75%

Heidrich et al. (2014)

0.9

1.0
2.6 L/LR/d

60%
76%

Gil-Carrera et al. (2013)

Raw municipal

6 g/L 4.1 kg/m3 2 g/L

6 g/L

COD, Chemical oxygen demand.

Carbon felt/nickel coated polyester cloth

38

CHAPTER 2 Application of bioelectrochemical systems

3
6 hours), with 2000 mg/Lr/d of OLR. In spite of domestic wastewater intended applications in the MEC, it was also observed producing less hydrogen yield during the early stages. Nevertheless, improvements in the MEC configurations and advanced process control suggest that better hydrogen yield is possible using domestic wastewater (Escapa et al., 2012).

2.7 Industrial wastewater The organic and inorganic content of wastewaters differs based on the type of industries. Methanol-rich effluent and food wastewater are assessed in this section. Methanol-rich wastewater contained a high methanol concentration of 1540 mg/L with less biodegradable matter while the food wastewater contained high levels of carbohydrates at around 2000 mg/L with lower amounts of acetate. It was observed that the physicochemical characteristics of the wastewater affect the catalyst performance, power densities, and total chemical oxygen demand (TCOD) removals. High levels of TCOD removals (90%) were observed in MEC treated with methanol wastewater, which indicates both the high biodegradability and efficient removal rates. A substantial amount of biogas (0.4 m3/m3/d) was captured in MEC using cathode catalysts and using methanol wastewater (Tenca et al., 2013). Baeza et al. operated 130 L of MEC reactor to treat raw glycerol and urban wastewaters. In their study, he achieved the hydrogen production rate of 4 L/d with 95% gas purity, 82% gas recovery, and 121% of total energy recovery with respect to the electrical input. The organic matter removal efficiency was approximately 25% for a hydraulic retention time of 2 days with an OLR of 0.25 gCOD/L/d. It should be possible to achieve removal efficiencies around 75% with OLRs lower than 0.05 gCOD/L/d. These results are promising and represent an important step toward the industrial implementation of these systems (Baeza et al., 2017).

2.8 Fermentation effluent The fermentation effluent could be a good source for hydrogen production. It includes byproducts such as acetic acids, butyric acids, propionic acids, lactic acids, formic acids, and ethanol. Rivera et al. studied the organic matter consumption and hydrogen production rate was evaluated in a two-chamber MEC. The robustness and the performance of the MEC were evaluated using either an anionic anion exchange membrane (AEM) or cationic exchange membrane (CEM) fed with VFAs. The study proved to attain the highest production rates of 81 mL/L/d, which were obtained with 550 mV, and 85% COD consumption (Rivera et al., 2015). Single-chambered MEC fed with 22.8 kg COD/m3/d OLR of ethanol fermentation effluent gave an overall hydrogen recovery of 96% and

2.9 Nutrient and heavy metals removals in MEC

production rate of 2.11 m3 H2/m3/d, corresponding to electrical energy efficiency of 287% (Lu et al., 2009). Wu et al. studied the anaerobic batch reactor (ABR) coupled with MEC to strengthen the H2 production. He conducted orthogonal experiments to operate ABR with COD of 4600 mg/L, HRT 2 days, and C/N ratio of 44 to achieve the high acetic acid accumulated effluent. MEC with nickel catalyst and carbon electrode, and 0.6 V fed with the ABR effluent, gives hydrogen production of 1.33 m3H2/m3/d with 99% COD removal (Wu et al., 2013).

2.9 Nutrient and heavy metals removals in MEC MEC had recently received great attention because it can simultaneously recover nutrients from wastewater and add value to resource recovery, which makes them more attractive over other nutrient-recovery technologies. Li et al. studied integrated microbial desalination cell (MDC)
microbial electrolysis cell (MEC) to perform energy-positive nutrient removal. The hybrid MDC
MEC achieved 10.0267 kW h/m3 energy balance with the total removal of 95.1% nitrogen, 63.7% saline, and 99.5% lead within 48 hours (Li et al., 2017). Sciarria et al. used the MEC to remove the nutrients in the anodic region and simultaneous H2 production in the cathodic region. The researchers reported the enhanced phosphorous recovery from high solid content digestate by coupling microbial electrochemical technologies (MET) with a crystallization process. Experiments were conducted using the digestate from an anaerobic digester, which was used as the feeding substrate for a microbial fuel cell (MFC) and an MEC to produce electricity and H2, promoting phosphorous removal. MECs were equipped with low-cost stainless steel mesh (SSM) cathodes to be compared with a benchmark platinum electrode, which were used to reduce MET cost. MEC systems (applied with 11.07 V) achieved 1.90 6 0.04 LH2/L/d using SSM and 2.02 6 0.03 LH2/L/d using a platinum cathode. The treatment led to phosphate (PO32 4 ) reduction between 21% and 30%, and COD between 27% and 44% of the digestate. The results obtained showed that the combination of the system proposed in this work with a biogas plant could be a low-cost solution for the recovery of nutrients and for hydrogen or electricity generation from digestate (Sciarria et al., 2019). Zhan et al. developed a new approach to achieve an autotrophic nitrogen removal from ammonium at a low applied voltage in a single-compartment, three-dimensional MEC. The MEC consisted of anodic and cathodic electrodes, on which nitrifying and denitrifying biofilms, respectively, were attached. Nitrogen removal can be enhanced at an applied voltage in the MEC. The nitrogen removal efficiency gradually increased from 70.3% to 92.6% with the increase of applied voltage from 0.2 to 0.4 V, as well as the maximum current, was varied from 4.4 to 14 mA. The corresponding coulombic efficiency (CE) also increased from 82% to 94.4%, indicating that the increasing applied voltage could enhance electron extraction from ammonium during its oxidative removal. The

39

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CHAPTER 2 Application of bioelectrochemical systems

dissolved oxygen (DO) was found to be a critical factor that affected the nitrogen removal in this MEC system. These results demonstrated that the MEC process was applicable to achieve autotrophic nitrogen removal from wastewater containing ammonium (Zhan et al., 2012). Colantonio et al. studied lead (Pb21) removal by the process called reduction and precipitation at the cathodic region using labscale MEC. The study investigated independent Pb21 removal using an anion exchange membrane at various voltage ranges 0.3, 0.6, and 0.9 V including open circuit. A considerable amount of metallic Pb21 (0. 10 6 0.02 mg) was found in the graphite fiber anode. Sufficient lead removal of 0.1
2.5 mg/L was observed at anode potential 20.15 to 20.33 V versus standard hydrogen electrode (SHE). However, the significance of exoelectrogens remained intact as the suppression of exoelectrogens with ethanol resulted in no Pb21 removal (Colantonio and Kim, 2016). In another study, Luo et al. developed a selective MEC to recover Cu21 and 21 Ni ions from wastewater. The study reported 97% 2 99% of Cu21 removal and 74.2% of Ni21 removal at an initial heavy metal concentration of 500 mg/L for both using maximum power densities increased from 3.1 6 0.5 to 5.4 6 0.6 W/m3 (Luo et al., 2015). Similarly, cathode materials for Ni21 reduction in MEC using electroplating effluents were assessed by Cai et al. (2016). The authors compared different cathodic materials such as copper sheet (CS), carbon cloth (CC), SSM, and graphite plate (GP) and evaluated the electrochemical performance, energy consumption rate, and the Ni21 removal. Results revealed that CS cathode showed the best performance with the cathode efficiency, energy consumption, and Ni21 removal of 148.56%, 40%, and 0.61 kW h/kg, respectively. Successful sulfide removal with the simultaneous biogas production from carbon dioxide using MFC
MECs coupled system was reported (Jiang et al., 2014). MEC was established at the applied voltage of 0.7 V to remove 72% sulfide and 57% methane formation within 70 hours of operation. Molecular characterization of biocathode determined the presence of the electro-active microbial community such as Methanobacterium palustre, Methanobrevibacter arboriphilus, and Methanocorpusculum parvum. The two-sided cathode MEC configuration shows the possibility to increase the efficiency of a three-chambered MEC aimed to nutrients and energy recovery. The performance of the dual-chambered MEC has been enhanced by anaerobic biological fermentation of urine to remove a higher COR removal rate. The influence of the prefermentation on the anode capacity was achieved for high current density of 218 6 6 mA/m2 and CE of 17% with the COD removal of 0.14 6 0.02 g/L/d (Barbosa et al., 2019).

2.10 Integrated MEC approach Among the various biological processes, DF for H 2 production is highly favorable due to the low energy requirement and low cost. Combining DF is

2.10 Integrated MEC approach

a promising strategy to recover higher H 2 yield with the possibility of obtaining a self-sufficient system. The integrated system consists of fermentation bioreactors, MEC/MFC. The number of the MEC/MFC attached to the DF bioreactor may vary between one to three. Few studies have utilized MFC/MEC to process the residual organic-containing effluent generated from the DF. The effluent is rich in VFAs to improve substrate utilization and energy recovery. Varanasi et al. investigated the integration of DF and MEC using water hyacinth as a substrate. Single chamber tubular MEC was fabricated to give overall hydrogen yield of 67.69 L H2/kg COD, COD removal of 70.33%, and energy recovery of 46% (Varanasi and Das, 2020). Due to huge volume availability, palm oil mill effluent (POME) is a proper feedstock for thermophilic H 2 fermentation with a high hydrogen production rate (about 16.9 mmol-H 2/L/h) (Krishnan et al., 2019). Undiluted raw pome gave an H 2 yield and H2 production rate of 236 mL-H 2/gCOD and 7.81 L/L/d with an 86% COD removal rate. The overall energy yield of 4.48 kJ/gCOD was achieved in the two-stage biohydrogen processes (Khongkliang et al., 2019). Other integrated approaches in MEC include dye-sensitized solar cell (DSSC)-powered MEC, biophoto electrochemical cell (BPEC), MFC
MEC coupled system, microbial reverse-electrodialysis electrolysis cells, microbial saline-wastewater electrolysis cell, and microbial electrolysis desalination and chemical production cell. Anaerobic digestion is the common degradation of complex organic matter from the residual of palm oil processes. Because the process involves high organic contents, it requires ample time to adapt and further break down and sustain organic matter (Krishnan et al., 2017a). In the process of degrading POME, the sequence of reactions is similar to other organic-anaerobic processes such as hydrolysis, acidogenesis (including acetogenesis), and methanogenesis, which are suitable for improving the existing treatment facility. Methanogenesis is the ratelimiting step in the anaerobic digestion of POME, whereas the conventional anaerobic digesters require large reactors and long retention time to ensure complete digestion of treated influent. Nonetheless, high-rate anaerobic bioreactors have been proposed (Krishnan et al., 2017b) to reduce reactor volume, shorter retention time as well as capture methane gas for utilization. To optimize the production rate of methane and increase the potential of other types of biogas (i.e., hydrogen), a two-stage anaerobic digestion is proposed. Hydrogen production coupled with methane production using MEC has been reported in the previous studies. Zhen et al. studied underlying mechanisms for methane bioelectrosynthesis, a two-chamber MECs containing a carbon biocathode. Considerable methane yield was achieved at a poised potential of 0.9 V (vs Ag/AgCl), reaching 2.30 6 0.34 mL after 5 hours of operation with a faradaic efficiency of 24.2%. Utilizing pome as substrate, an integrated approach CSTR and MEC was established for continuous hydrogen and methane production. The system obtained the methane yield and production rate of 290 mL CH4/gCOD and 2700 mL CH4/L at 8 days of HRT.

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CHAPTER 2 Application of bioelectrochemical systems

2.11 Conclusions As a novel alternative hydrogen production process, MECs have attracted attention because of their high hydrogen conversion efficiency, low energy requirements, and applicability to many organic substrates. With recent advancements and major technological breakthroughs made over the past decade, many inherent challenges continue to obstruct MECs large-scale and also in real-world applications. Detailed understanding of the MEC and crucial factors governing MEC is needed to be regulated. These include exoelectrogens, cathode and anode materials, substrate, voltage, inoculum, temperature, electrolyte, HRT, and OLR. Further research will be required to understand the impact of various operational parameters and optimize the MEC’s hydrogen recovery. Continuous progress in MEC results in the increased production rate and low operation costs and make the MEC system more sustainable.

Acknowledgments This research is financially supported by PDRU Grant- Vot No. Q.J130000.21A2.04E53, The Hitachi Global Foundation, MRUN R.J130000.7805.4L886, and Research University Grant Scheme (Vote Q.J130000.2609.09J40) which are gratefully acknowledged.

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17213. ` ., Guerrero, J., Ruiz, Y., Guisasola, A., 2017. BioelectrocheBaeza, J.A., Mart´ınez-Miro´, A mical hydrogen production from urban wastewater on a pilot scale. J. Power Sources 356, 500
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527. Mishra, P., Krishnan, S., Rana, S., Singh, L., Sakinah, M., Ab Wahid, Z., 2019. Outlook of fermentative hydrogen production techniques: an overview of dark, photo and integrated dark-photo fermentative approach to biomass. Energy Strategy Rev. 24, 27
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11123. Zhan, G., Zhang, L., Li, D., Su, W., Tao, Y., Qian, J., 2012. Autotrophic nitrogen removal from ammonium at low applied voltage in a single-compartment microbial electrolysis cell. Bioresour. Technol. 116, 271
277.

CHAPTER

Nutrient removal and recovery in bioelectrochemical systems

3

Aryama Raychaudhuri and Manaswini Behera School of Infrastructure, Indian Institute of Technology Bhubaneswar, Bhubaneswar, India

3.1 Introduction Macro- and micronutrients are essential for living organisms for growth and reproduction. As plants obtain these nutrients from soil, chemical fertilizers are used to maintain the nutrient reserve in agricultural land. Fertilizer production process encompasses the transformation of a large amount of atmospheric nitrogen (120
160 Mt per year) into reactive ammonia by the Haber
Bosch process (Arredondo et al., 2015). A substantial quantity of unused nutrients (nitrogen and phosphorus) eventually enters the natural water bodies as agricultural runoff. Excretory products of humans and animals are also rich in nutrients that enter the environment as sewage and manure. Inefficient treatment of domestic and industrial wastewater (such as swine wastewater, landfill leachate, coke wastewater, dairy manure, urine waste, anaerobic digestor liquor, and beverage wastewater) may lead to nutrient pollution in surface and subsurface water resources (Kumar and Pal, 2015). Nitrate toxicity of the subsurface water has been known to cause blue baby syndrome in infants due to the consumption of water with excess nitrate content (Nancharaiah and Venugopalan, 2011). Nitrogen and phosphorus accumulation in surface water induce the development of algae and other aquatic plants resulting in depletion of dissolved oxygen (DO) in the water body and eventually causing fish kill. Through affecting fisheries and tourism, marine eutrophication has a detrimental effect on food security, economy, ecosystem health, and the overall environment (Ngatia et al., 2019). The removal and recovery of nutrients are of utmost concern given to the potential threat to water bodies, limited mineral resources, and high expense of fertilizer production (Rittmann et al., 2011). Ammonia, the primary form of nitrogen in wastewater, is removed through two sequential processes: conversion of ammonia to nitrate by aerobic nitrification and conversion of nitrate to N2 gas by anoxic denitrification. This process is energynegative and carbon-intensive, which requires a large amount of aeration, generating a significant amount of sludge (Yu et al., 2011; Li et al., 2016a). Additional biological and physicochemical technologies to remove ammonia from wastewater includes anaerobic ammonium oxidation (ANAMMOX), ammonia stripping, and ion exchange

Delivering Low-Carbon Biofuels with Bioproduct Recovery. DOI: https://doi.org/10.1016/B978-0-12-821841-9.00001-3 © 2021 Elsevier Inc. All rights reserved.

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(Ahn, 2006; Gupta et al., 2015; Liu et al., 2016; Christiaens et al., 2019). However, these treatment processes incur high cost, making them an unsustainable solution. Phosphorus can be removed by enhanced biological phosphorous removal (Barnard, 1975), membrane technologies (Zhang et al., 2013; Xie et al., 2014), and chemical precipitation (struvite) (Jaffer et al., 2002; El Diwani et al., 2007). Struvite can remove both ammonia and phosphate from wastewater, but the cost associated with the process is high, and the application of struvite is not yet widespread in agriculture (Le Corre et al., 2009; Zhang et al., 2014a). Microalgae can consume nitrogen and phosphorus from wastewater and generate biomass from which value-added products can be extracted. However, separation of algae from treated wastewater can be challenging (Cai et al., 2013). Since both nitrogen and phosphorus are essential components of fertilizer, the recovery of these nutrients and conversion into a reusable form can efficiently treat wastewater, reduce energy consumption, and avoid exhaustion of natural resources, thus establishing a sustainable energywater-nutrient nexus (Gao et al., 2018). In the last decade, bioelectrochemical systems (BES) have emerged as a promising technology for the removal of contaminants and recovery of valuable materials with direct electricity generation. Electrochemical cells that take advantage of microorganisms to catalyze oxidation and reduction reactions typically fall under BES (Nancharaiah et al., 2016). The two main representatives of BES includes microbial fuel cell (MFC) and microbial electrolysis cell (MEC). In MFC, electrochemically active bacteria oxidize the organic matter and deliver the electrons to the anode electrode while releasing protons in the solution. The electron travels through an external circuit to cathode reducing oxygen, and to maintain charge neutrality, protons migrate to the cathode through the cation exchange membrane (CEM) (Logan et al., 2006). Whereas in MEC, hydrogen is produced in the cathode under the influence of externally supplied voltage and a biologically assisted condition (Kadier et al., 2016). In addition to energy recovery with simultaneous wastewater treatment, other applications of BES include water desalination (microbial desalination cell, MDC) (Roy and Pandit, 2019; Brastad and He, 2013), synthesis of chemicals (microbial electrosynthesis) (Rabaey and Rozendal, 2010), CO2 sequestration using algae (microbial carbon capture cell, MCC) (Jadhav et al., 2017; Das et al., 2019), bioremediation (benthic/sediment microbial fuel cell, BMFC/SMFC) (Liu et al., 2015), metal (Wang and Ren, 2014), and nutrient recovery (Kelly and He, 2014) (Fig. 3.1). While ammonia assimilation and biochemical denitrification (coupled with nitrification) is utilized in BES for the removal of nitrogen, the current driven ammonium ion migration from anode to cathode is of utmost importance for the advancement of the recovery process (Zhang et al., 2014a; Nancharaiah et al., 2016). Efficient phosphorus removal and recovery in BES generally involve struvite precipitation, which relies on high cathodic pH (Ichihashi and Hirooka, 2012). Thus, BES has possible benefits relative to current technologies in terms of energy efficiency and low carbon footprint.

3.2 Nitrogen removal and recovery

FIGURE 3.1 Schematic diagram of four typical dual-chamber BES. (A) Electricity generation in MFC; (B) biohydrogen recovery in MEC; (C) desalination of saline water in MDC; and (D) recovery of nutrients in MNRC. BES, Bioelectrochemical system; MDC, microbial desalination cell; MEC, microbial electrolysis cell; MFC, microbial fuel cell; MNRC, microbial nutrient recovery cell.

3.2 Nitrogen removal and recovery The Haber
Bosch process is the predominant industrial method to fix atmospheric nitrogen to ammonia. In this conversion process, nitrogen combines with hydrogen at high temperature (400 C
600 C) and pressure (20
40 MPa) in the presence of an iron catalyst generating approximately 1.6 3 1011 kg of ammonia annually. This process consumes nearly 1%
2% of the world’s annual energy production. Most of the ammonia is used in chemical fertilizer, and a substantial amount of unutilized ammonia ends up in the environment, causing pollution (Arredondo et al., 2015; Nancharaiah et al., 2016). The conventional nitrogen removal process is energy

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extensive, which converts ammonia into nitrogen gas. A significant amount of energy is invested in releasing fixed nitrogen in the atmosphere, and an even more energyintensive chemical process is used to recapture atmospheric nitrogen into reactive ammonium compounds (Kim et al., 2015). The recovery of ammonium nitrogen from wastewater is thus deemed a more sustainable solution than removal of it. BES offers several advantages as an alternative to traditional nitrogen removal methods. This section provides a summary of nitrogen removal and recovery mechanisms in BESs.

3.2.1 Issues related to conventional technologies Nitrogen is often found in the form of ammonium in wastewater, and its removal in wastewater treatment plants occurs mostly through two sequential processes. In the first step, autotrophic nitrification under aerobic condition, ammonium ion (NH1 4 ) is oxidized by ammonia-oxidizing bacteria (AOB, e.g., Nitrosomonas) to nitrite (NO2 2 ), and nitrite-oxidizing bacteria eventually oxidize the nitrate (NOB, e.g., Nitrobacter) to nitrate (NO2 3 Þ. As 1 g of ammonium nitrogen is completely oxidized to nitrate, 4.7 g of O2 is consumed. During the second step, heterotrophic denitrification under anoxic condition, denitrifying bacteria (e.g., Paracoccus denitrificans), oxidizes organic matter to CO2 and reduces nitrate to N2 gas which is released to the atmosphere (Arredondo et al., 2015; Nancharaiah et al., 2016). Generally, autotrophic bacteria (AOB and NOB) perform well at high DO level, low organic loading, and long sludge retention time, whereas most of the heterotrophic denitrifiers require ample organic loading. So, the conventional biological nitrogen removal (BNR) processes have several disadvantages such as high energy requirement for aeration, large space requirement, addition of alkalinity to regulate pH, requirement of electron donor (methanol) for denitrification, and high sludge generation (Li et al., 2016a; Zhang and He, 2012a). In the early 1990s, an energy-efficient alternative to BNR was discovered, which is known as ANAMMOX. In anaerobic condition, ANAMMOX bacteria (Planktomycetes such as Candidatus Kuenenia stuttgartiensis) convert NH1 4 to N2 using NO2 as a terminal electron acceptor (Kuenen, 2008). This process is advan2 tageous over BNR because ANAMMOX needs aeration for only partial oxidation of ammonium to nitrite (60% less aeration compared to BNR), and involvement of autotrophic organisms eliminates the requirement of electron donor (90% reduction in sludge generation) (Lackner et al., 2014). The ANAMMOX process can be described by Eqs. (3.1) and (3.2) (Arredondo et al., 2015). 2 1 NH1 4 1 1:5O2 -NO2 1 H2 O 1 2H

NH1 4

1 NO2 2 -N2

1 2H2 O

(3.1) (3.2)

The ANAMMOX needs the availability of nitrite, which is usually not present in wastewater. An additional pretreatment phase is therefore necessary to partially oxidize the ammonium to nitrite. So, ANAMMOX is coupled with the SHARON (single reactor system for high-rate ammonium removal over nitrite) process. In the

3.2 Nitrogen removal and recovery

SHARON process, controlled oxidation of ammonium up to nitrite takes place using the specific growth rates of AOB and NOB and by regulating the ammonium loading rate (Van Dongen et al., 2001). An alternative approach is called CANON (completely autotrophic nitrogen removal over nitrite) process where AOB and ANAMMOX bacteria symbiotically coexist for the simultaneous aerobic and anaerobic oxidation of ammonium in a single reactor (Jetten et al., 2001; Third et al., 2001). Recently, a process has been developed which can generate energy with simultaneous removal of nitrogen. In the CANDO (coupled aerobic
anoxic nitrous decomposition operation) process, the ammonium is converted to nitrite by SHARON, then the nitrite is partially reduced to nitrous oxide and NO2 further decomposed to N2 and O2 producing energy (Scherson et al., 2013). The ANAMMOX, the SHARON, CANON, and CANDO processes require less energy and generate relatively less sludge in comparison to BNR, as well as offer energy recovery, but it cannot achieve nitrogen recovery.

3.2.2 Nitrogen removal in bioelectrochemical system BES holds potential as a sustainable solution to wastewater treatment with simultaneous recovery of energy and other value-added products. For efficient wastewater treatment, both organic pollutants as well as nutrients removal is essential. As nitrogen removal necessitates aerobic conditions, the anaerobic environment in the anode of BES does not allow efficient ammonium nitrogen removal from wastewater. In the last decade, a variety of scientific studies have concentrated on establishing an effective strategy for the removal and recovery of nitrogen in BES. Few studies have reported the effect of nitrogenous compounds on the performance of BES in terms of electricity generation. Biological nitrification has been examined in the biocathode of an MFC by adding ammonium to the catholyte along with nitrifying bacteria. Improvement in the performance of the MFC was observed primarily due to the lower pH (8.8 27) of the catholyte. The protons produced by biological nitrification may buffer the elevated pH of the catholyte owing to the reduction of oxygen (You et al., 2009). Further research verified the fact that nitrification in the cathode of BES can consume alkalinity and reduce the pH (Virdis et al., 2010). However, nitrification activity can also have a negative impact on BES performance. In a study, performance of the MFC was examined with different ammonium concentrations (30
100 mg/L) in the cathode. A decrease in cathodic potential was observed with an increase in ammonium concentration. The nitrifying biomass consumed more oxygen with higher ammonium levels resulting in inadequate DO concentration for cathodic reduction reaction (Ryu et al., 2013). Denitrification in the anode can also deteriorate the power output of the cell. It was reported that the presence of 4
8 mM nitrate in an aircathode MFC decreased the electricity generation, perhaps owing to the competition for the substrate (electron donor) between the anode-respiring exoelectrogens and denitrifying bacteria (Sukkasem et al., 2008). Therefore, it is imperative to understand the nitrogen-related mechanisms in BES, which will not only control its impact on BES performance but also remove the pollutants and/or recover valuable nutrients. The basic mechanism of ammonia removal is presented in Fig. 3.2.

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FIGURE 3.2 Depiction of different removal mechanism of ammonia in BES. ANAMMOX, Anaerobic ammonium oxidation; BES, bioelectrochemical system; PBES, photo-bioelectrochemical system; SND, simultaneous nitrification and denitrification.

The processes reported for ammonia removal in BES comprise of four main mechanisms. The basis of the first mechanism is the transport of ammonia/ammonium through the CEM via either diffusion or current driven migration (Kim et al., 2008). Ammonia (NH3) loss can occur by passive diffusion through the membrane driven by the concentration gradient. NH1 4 concentration in wastewater is generally much higher than protons [NH1 4 concentration in urine is .4 g/L (Maurer et al., 2006) and in domestic wastewater 0.04 g/L (Arredondo et al., 2015)]; therefore, ammonium ions may migrate from anode to cathode chamber to maintain electroneutrality. Due to the elevated pH of the catholyte the equilibrium shifts from ammonium to ammonia gas. This NH1 4 /NH3 pair was used to regulate the pH of electrolyte in BES (Cord-Ruwisch et al., 2011). The ammonium ions were added in the anode of a dual-chambered MFC, which prevented anolyte acidification and enabled the enrichment of electroactive anodic biofilm. The ammonium ions migrated to the cathode compartment across the membrane, where it was converted to ammonia gas because of the prevailing alkaline pH

3.2 Nitrogen removal and recovery

(around 9.5), which thereafter stripped from the catholyte and reintroduced to the anolyte to maintain the pH (around 7) for the sustained current generation. This concept was further employed in MEC to avoid the energy expenditure on stripping (Cheng et al., 2013), where the hydrogen gas formed at the cathode assisted the volatilization of dissolved ammonia and the ammonia gas was escaped from the cathodic compartment and enter the anodic compartment. In these studies, the ammonium migration was used as a proton shuttle to maintain the operations in BES, and the findings have prompted further research on the recovery of ammonium nitrogen in BES. The second mechanism for the removal of nitrogen in BES is the denitrification of nitrate to nitrogen gas in the cathode chamber. For this process to occur, first ammonia should be transformed into nitrate by aeration (nitrification). Nitrate can serve as a terminal electron acceptor and generate a positive electric potential in this process utilizing organic compounds (such as acetate) as an electron donor, as described in Eqs. (3.3) and (3.4) (Kelly and He, 2014). 0

CH3 COO2 1 2H2 O-2CO2 1 7H1 1 8e2 E0 5 2 0:28V versus SHE 2NO2 3

2

1

0

1 10e 1 12H -N2 1 6H2 O E0 51 0:70V versus SHE

(3.3) (3.4)

Conventional denitrification is achieved by heterotrophic denitrifying bacteria, whereas bioelectrochemical denitrification is accomplished by autotrophic denitrifying bacteria, which are capable of receiving electrons from a solid electron donor (such as cathode electrode). Therefore, supply of additional carbon source is not required for nitrate reduction (Kelly and He, 2014). This phenomenon was demonstrated for the first time by Geobacter species as a pure culture of Geobacter metallireducens has been found to be capable of reducing nitrate to nitrite with graphite electrode as the sole electron donor (Gregory et al., 2004). Another study demonstrated complete denitrification at the cathode of an MFC without external energy supply by utilizing electrons supplied through the cathode electrode provided via microbial oxidation of acetate in the anode (Clauwaert et al., 2007). Moreover, ammonia can be anaerobically oxidized to N2 gas via ANAMMOX using nitrite as an electron acceptor. This process will produce a positive electrical potential under a standard condition theoretically; however, this thermodynamically favored method has a rather slow kinetics and ANAMMOX bacteria is difficult to cultivate as well (He et al., 2009). A novel integrated MEC-MFC system was explored, which accelerates the ANAMMOX process while alleviating the dependency on slowgrowing ANAMMOX bacteria. Batch experiments found that the MEC-MFC system eliminated 85% of total nitrogen (TN) within 10 days, while ANAMMOX removed only 62% of TN. No external carbon source was added as wastewater energy was efficiently utilized, thereby ensuring self-sufficient nitrogen removal (Li et al., 2016b). In 2 the ANAMMOX process, 1 mol of NH1 4 2 N is oxidized with 1.32 mol of NO2 2 N, 2 whereas 0.26 mol of NO3 2 N was formed as a byproduct. Therefore, the theoretical maximum removal capacity of TN in ANAMMOX is only 89%, which does not comply with the discharge limit of NO2 3 2 N (Zhang et al., 2019; Sliekers et al., 2002;

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Cao et al., 2019). To address this issue, ANAMMOX process was combined with a denitrifying MFC system. In the cathode chamber of the MFC, NH1 4 2 N and 2 NO2 2 N were consumed by ANAMMOX bacteria, whereas NO 2 N and the pro2 2 duced NO2 2 N acts as electron acceptor and converted to nitrogen gas by denitrifying 3 bacteria. Microbial community analysis revealed the presence of Candidatus Brocadia sinica as the main ANAMMOX bacteria and Rhodopseudomonas palustris as the autotrophic denitrifier, who can accept electrons from cathode electrode. The study demonstrated coexistence of these two kinds of microbial community in the BES for the effective removal of nitrogenous compounds from wastewater (Li et al., 2015). Further, ANAMMOX bacteria was used as a biological cathode catalyst in MDC to investigate the possibility of simultaneous wastewater treatment in the anode chamber, desalination of salt water, and removal of nitrogenous compounds in the cathode chamber. A maximum power of 0.092 W/m3 was obtained with more than 90% ammonium removal. Moreover, treated effluent from the anode was introduced to the cathode for the removal of nitrogenous compounds in the wastewater stream. A decline in performance was observed because of the growth heterotrophic organisms at high organic concentration. It was concluded that the system is particularly suitable for wastewater with low C/N ratio, and ANAMMOX bacteria perform better at high ammonium concentration and low DO level (Kokabian et al., 2018). A recent study demonstrated a BES containing membrane-aerated nitritation-ANAMMOX in its cathode. The BES achieved the maximum TN and chemical oxygen demand (COD) removal efficiency of 95% and 98%, respectively, at hydraulic retention time (HRT) of 60 hours (Yang et al., 2017). ANAMMOX bacteria have also been tested for their ability to transfer electrons to the electrode. It was used as an anode inculum in an air-cathode MFC to drive nitrate reduction in the cathode with simultaneous electricity generation without an additional carbon source. A TN removal efficiency of 96.3% was achieved in the system. Microbial community analysis revealed the presence of Candidatus Kuenenia as the predominant ANAMMOX community. Also, the presence of Cytochrome c (containing Heme c) indicated the possibility of extracellular electron transfer by the ANAMMOX bacteria (Zhang et al., 2020). Although these studies show promising results for nitrogen removal, nitrogen recovery is not possible. The third mechanism that has been proposed occurs at the anode, where ammonium is transformed directly by microorganisms into N2 gas. Ammonium can be utilized as a fuel in the anode to generate electricity or it can act as a substrate for nitrifiers to produce organic compounds for heterotrophs, yet there is also no convincing justification for this (He et al., 2009). The fourth mechanism involves ammonium uptake in the anode or cathode compartment of BES by microbial biomass for their growth and development. The main mechanism of ammonia removal in BES is depicted in Fig. 3.3.

3.2.2.1 Reactor configuration for bioelectrochemical nitrogen transformation As discussed earlier, ammonia has found to be the main nitrogenous compound present in wastewater. The process of nitrate reduction (denitrification) can be

3.2 Nitrogen removal and recovery

FIGURE 3.3 Schematic representation of different nitrogen removal mechanisms in BES. (1) Incorporation of ammonium in biomass at anode; (2) passive diffusion of ammonia through membrane; (3) electromigration of ammonium ion through membrane; (4) conversion of ammonium to ammonia; (5) conversion of aqueous ammonium to ammonia gas due to elevated pH of the catholyte; (6) nitrification of ammonia; (7) denitrification of nitrate by SND; (8) shortcut nitrification; (9) autotrophic denitrification or ANAMMOX; and (10) ammonium uptake by algal biomass in a PBES. ANAMMOX, Anaerobic ammonium oxidation; BES, bioelectrochemical system; PBES, photobioelectrochemical system; SND, simultaneous nitrification and denitrification.

realized in the cathode of the BES but for the removal of TN from real wastewater first ammonia should be converted to nitrate for subsequent bioelectrochemical denitrification. Simultaneous oxidation of organics in the anode and reduction of nitrate in the cathode was successfully demonstrated from two different fluid streams wherein the electrons generated at the anode were utilized to faciliate the nitrate reduction at the cathode (An et al., 2016; Al-Mamun et al., 2017). Complete nitrogen removal was demonstrated in BES for the first time through incorporation of an aerobic process. In this system, simultaneous power generation, carbon, and nitrogen removal from wastewater was achieved by using a denitrifying MFC coupled with a biofilm-based aerobic nitrifying bioreactor (Virdis et al., 2008). The wastewater was first fed to the anode chamber of the MFC where organic compounds were oxidized to generate electrons, which aided the reduction of nitrate in the cathode; then the anode effluent containing ammonia was pumped into the external reactor where ammonia was aerobically oxidized to nitrate; and finally the nitrified liquid entered the cathode chamber of the

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MFC where nitrate was reduced to nitrogen gas. This system attained a maximum nitrogen removal rate and power output of 0.41 kgNO3-N/m3 per day and 34.6 W/m3, respectively. Though complete TN removal was expected, elevated ammonia concentration was found in the cathode largely because of ammonia crossover through CEM from anode to cathode chamber. This issue was successfully resolved in their subsequent design by the incorporation of the nitrification stage into the cathode chamber, thereby facilitating simultaneous nitrification and denitrification (SND) in a single compartment (Virdis et al., 2010). The synthetic wastewater was pumped into the anode, and the effluent was injected into the cathode through the loop connection, while oxygenation of the catholyte was done in the recirculation line. Parameters such as the DO level and C/N ratio were optimized and 94.1% nitrogen removal efficiency was attained in the MFC. It was assumed that, in the presence of oxygen denitrifiers can survive in the cathode benefiting from the biofilm-electrode framework that could establish a microanoxic environment. Additional examination of the cathode biofilm stratification showed that the nitrifying bacteria had developed in the outer layer of the biofilm and the putative denitrifying microbes had inhabited the inner layer, supporting the viability of the SND in the MFC cathode (Virdis et al., 2011). Intermittent aeration was applied in the cathode of a dual-chambered MFC to establish SND. The presence of both nitrifiers (Nitrosomonas sp.) and heterotrophic denitrifiers (Alcaligenaceae and Comamonadaceae) confirmed the feasibility of SND (Sotres et al., 2016). To reduce the cost of the BES and improve nutrient removal, rotating biological contractor (RBC) was integrated with membrane-less MFC. SND was observed in the cathode, as the cathode disk rotation exposes it partly to atmospheric oxygen, thus creating an oxygen gradient in the system (Sayess et al., 2013). Due to the fact that SND requires meticulous control of DO, MFC systems were equipped with separate nitrifying (aerobic) and denitrifying (anoxic) cathodes. The system consisted of two MFCs, first having dual aerobic biocathodes (O-MFC) and the second containing dual anoxic biocathodes (A-MFC). The synthetic wastewater was injected into the anodes of the two MFCs separately (where the COD was primarily removed); the effluents were collectively directed to the cathodes of O-MFC (where ammonium was oxidized to nitrate), whose effluents were then transferred into the cathodes of A-MFC (where nitrate was electrochemically denitrified to nitogen gas). Maximum COD, TN, and NH1 4 2 N removals of 98.8%, 97.3%, and 97.4%, respectively, was achieved (Xie et al., 2011). This reactor configuration has been simplified to a three-chambered dual-cathode MFC comprising aerobic and anoxic cathode on either side of the anode where nitrogen removal undergoes the similar oxic-anoxic process. However, the anode and aerobic cathode were separated by a CEM, whereas anode and the anoxic cathode were partitioned by an anion exchange membrane (AEM) (Zhang and He, 2012a). To increase the nitrogen removal rate, the MFC was further reconfigured to a continuously operated tubular MFC consisting of an inner anoxic cathode and an outer aerobic cathode, both sharing a common anode in between. Ammonium and nitrate removal were 96% and 89%, respectively

3.2 Nitrogen removal and recovery

(Zhang and He, 2012b). To reduce the cost of excess aeration and shorten the operation time of nitrification, this system was investigated for short-cut nitrification in the aerobic cathode, where ammonium is converted to nitrite instead of nitrite at low DO concentration (#3.5 mg/L). The effluent of this chamber was added to the anoxic cathode in batch mode where nitrate was reduced to nitrogen gas (using the electrons from the anode chamber) by autotrophic denitrification (Li et al., 2016a). Conventional MFC was combined with a membrane bioreactor (MFC-MBR (membrane bioreactor) system) to facilitate wastewater treatment and nitrogen removal. Conductive membrane modules (fabricated using stainlesssteel (SS) mesh and carbon felt) with attached biofilm was used as a cathode, where simultaneous filtration and bioelectrochemical denitrification took place. Microbial analysis of the cathodic biofilm revealed that AOBs and NOBs prevailed in the exterior biofilm executing ammonium oxidation, whereas denitrifying bacteria were more dominant adjacent to the cathode for trapping electrons (Zhang et al., 2014b). SND was also demonstrated in MFC with cathode chamber consisting of a membrane-aerated biofilm reactor, which ensured flexible control of the DO for efficient SND (Wu et al., 2017). Photo-bioelectrochemical systems (PBESs) provide a new approach for simultaneous nutrient removal and electricity generation by establishing synergistic collaborations between exoelectrogens and photosynthetic microorganisms (algae). Algae can assimilate nutrients for their growth with the help of various bacteria in the system. Incorporation of algae in the cathode chamber of the BES would aid the in-situ generation of oxygen through photosynthesis, which function as a terminal electron acceptor avoiding mechanical aeration (Jiang et al., 2019). Algal biomass was introduced in a sediment MFC and the system was able to remove carbon, nitrogen, and phosphorus simultaneously while generating electricity. TN removal of 87% was achieved in which algal biomass contributed to 75% and the reminder was removal through nitrification and denitrification (Zhang et al., 2011). Algal cathode MFC was investigated for the nutrient removal from landfill leachate. Maximum cell voltage was obtained with 5% leachate suggesting that proper dilution of leachate is necessary to reduce inhibitory level of ammonia and other toxic chemicals (Nguyen et al., 2017). A photoMFC was investigated for the cultivation of Spirulina in the cathode compartment while the anode compartment was fed with swine-farming wastewater. The nutrients were transported to the cathodic compartment by electromigration and diffusion and removed by algal uptake. The alkaline condition of the cathode did not appear to have a negative effect on the growth of Spirulina (Colombo et al., 2017). An algal biofilm MFC was established by integrating algal biofilm grown on a compressed glass wool, which constitute an interface between anode and cathode chamber. The COD and TN removal in the batch operation reached up to 80% and 96% while algal uptake was the main mechanism for nutrient removal (Yang et al., 2018). A novel in situ self-sustaining photo-MFC was developed, which was submerged in the eutrophic water for the nutrient removal. The anode and cathode chamber were separated by an AEM, a perforated plate, and a CEM

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in the specific order. Synthetic wastewater was fed in the anode camber and the effluent was mixed with Chlorella vulgaris concentrate and directed to the cathode chamber in a batch mode operation. The nutrient ions present in the surrounding eutrophic water were migrated to the respective chambers (NO2 3 2 N and 1 PO32 2 P in the anode chamber, NH 2 N in the cathode chamber) driven by 4 4 electric field generated by the MFC and recovered as microalgal biomass. Furthermore, oxygen supply by the photosynthetic microalgal biomass can significantly improve cathodic reduction reaction, thereby enhancing electricity generation (Jiang et al., 2019). To mimic the natural metabolic pattern of algae under day/night cycle, the algal-bacterial PBES was operated under alternating 12 hours light/12 hours dark pattern. The system can use oxygen (produced via photosynthesis) and nitrate/nitrite as electron acceptors under different operating conditions. The nitrogen removal can be realized via several approaches: nitrification using photosynthetic oxygen, bioelectrochemical denitrification using cathode as electron donor with nitrate as electron acceptor, heterotrophic denitrification using photosynthetically derived dissolved organics as carbon source, and algal absorption (Sun et al., 2019). In addition, constructed wetland-MFCs were also investigated as an eco-friendly and cost-effective system for nutrient removal from wastewater. The removal of nitrogen and phosphorous compounds were facilitated by plant uptake as well as natural aerobic and anaerobic processes involving autotrophic and heterotrophic microorganisms (Wang et al., 2019; Ge et al., 2020).

3.2.2.2 Groundwater remediation using bioelectrochemical system Nitrate must be removed from the groundwater since a significant part of the world’s population depends on groundwater as a drinking water source. Due to the unavailability of electron donors for heterotrophic denitrification, the treatment of nitrate contamination in groundwater is challenging (Tong and He, 2014). A cost-competitive, carbon-free BES was investigated for the removal of nitrate from groundwater. The potential of a denitrifying biocathode was regulated while using water as an anode electron donor, and the maximum nitrate removal efficiency (90%) was achieved at 2123 mV versus SHE. The utilization of water rather than acetate as an electron donor can reduce the operating cost as well as reduce the risk of groundwater contamination by acetate diffusion through membrane (Pous et al., 2015). A dual-chamber BES was investigated to remove nitrate from groundwater through autotrophic denitrification. The operation of the reactor was switched between MFC and MEC mode to enhance the efficiency of nitrate removal. Nitrate concentration in the effluent was low enough to meet the WHO guidelines for drinking water; however, the accumulation of N2O was observed (Molognoni et al., 2017). More efficient and complete removal of nitrate and nitrite was observed in a sequential two-stage BES employing controlled biocathodic denitrification systems (CBD). Unlike MFC, the electron supply to the cathode is not restricted by the oxidation rates at the anode in CBD, and the electron flux is provided by a power supply or a potentiostat. Nitrate was removed by autotrophic denitrification in the first stage and the accumulated N2O was

3.2 Nitrogen removal and recovery

removed in the second stage, which acted primarily as a polishing step (Cecconet et al., 2019). Groundwater remediation by an ex situ approach such as pump-andtreat is expensive and seldom requires the addition of chemicals and electron donors; therefore in-situ remediation without deterioration of the water quality is of prime concern. The principle of MDC was applied to develop a submerged microbial desalination
denitrification cell for the in situ removal of nitrate from groundwater. The potential difference between the anode and the cathode was 1 used to drive NO2 ions to the anode and cathode chamber through 3 and Na AEM and CEM, respectively. The nitrate-containing anode effluent was introduced to the cathode chamber where it was converted to nitrogen gas by autotrophic denitrification. The system achieved 90.5% nitrate removal within 12 hours and proved to be a promising application of BES for groundwater treatment (Zhang and Angelidaki, 2013). As nitrate would primarily enter the anode chamber, the system was further modified to accomplish nitrate removal in the anode of the BES. Under the influence of an external electric potential, NO2 3 ions were migrated to the anode compartment through AEM and then converted to nitrogen gas by heterotrophic denitrification in the presence of organic compounds (Tong and He, 2013). The system was advanced to prevent the leakage of anolyte to the groundwater by modifying the reactor configuration and providing multiple barriers. The BES consisted of two concentric membrane tubes: the inner tube was made of CEM and acted as anode, and the outer tube was made of AEM. The BES was submerged into the nitrate-containing groundwater, and the space between the two tubes created a concentrating chamber in which nitrate ions were migrated and accumulated under the influence of external electric current. Further treatment would be necessary to remove nitrate as the system accumulates nitrate rather than reducing it to nitrogen gas (Tong and He, 2014). Furthermore, energy demand related to in-situ and ex situ treatment of nitrate-contaminated groundwater in MFCs and CBDs were analyzed and compared under different reactor configurations/operations (Cecconet et al., 2018).

3.2.2.3 Influential operational parameters Various parameters such as C/N ratio, DO concentration, electrolyte pH, and electrode potential can significantly affect the performance of BES for the removal of nitrogen. The operating parameters such as HRT, recirculation rate, concentration/ dilution of the waste streams, and external resistance can also influence the system output (Kelly and He, 2014; Sun et al., 2016; Chen et al., 2016). Optimization of the operational parameters is crucial for stabilized operation. One of the major factors affecting the performance of BES is pH, which influences the microbial metabolism (Haddadi et al., 2013). The nitrogen removal rate doubled by maintaining the catholyte pH at 7.2 indicating that the pH split can affect the cathodic nitrate reduction (Clauwaert et al., 2009). It was demonstrated that ammonium/ammonia shuttle between the anode and cathode chamber can also help in maintaining the pH of the system (Cord-Ruwisch et al., 2011). An electrochemically active biofilm was developed, which acted as both bioanode and biocathode, and has been found to neutralize

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CHAPTER 3 Nutrient removal and recovery in bioelectrochemical

acidity produced from organics oxidation by alkalinity produced from cathodic denitrification (Cheng et al., 2012). DO concentration is another key influential factor, especially in BES with SND process. A study reported that maximum nitrogen removal by SND occurred at a DO level of 4.35 mg/L (Virdis et al., 2010); while other research has suggested that SND occurred at a DO level of 0.5 mg/L in a membrane-aerated MFC (Yu et al., 2011). Optimizing the DO concentration is imperative since low DO would promote accumulation of ammonia in the final effluent, whereas high DO would hinder denitrification, leading to accumulation of nitrate (Kelly and He, 2014). C/N ratio was also considered as an important operating parameter because it can affect by-product formation and nitrogen removal. It was observed that an increase in C/N ratio inhibited nitrate accumulation and assisted its removal (Huang et al., 2013). In a study, when C/N ratio was decreased from 2.5 to 0.97 the nitrate accumulation increased significantly (Zhou et al., 2007). However, a high C/N ratio may favor heterotrophic denitrification and hinder bioelectrochemical denitrification (Zhang and He, 2013). The anode and cathode potential play a key role in maintaining the bioactivity in the respective compartment. Regulating the cathode potential has been found beneficial for autotrophic denitrification. The external resistance may also influence the BES performance by affecting electron supply and current generation. The anodic biofilm structure may also be affected from higher external resistance. Reducing the external resistance will promote higher current flow to the cathode, which will prove bioelectrochemical denitrification. This process is demonstrated in a dual-cathode MFC in which the reduction in external resistance from 712 to 10 Ω improved nitrate removal from 52.1% to 66.4% (Zhang and He, 2012a).

3.2.2.4 Bacteriological approaches The microbial community grown on electrode surface can regulate the performance of the BES. Electroactive bacteria (EAB) grown on the anode are enriched with mixed anaerobic microbial communities or pure cultures with electrogenic properties exhibiting extracellular electron transfer such as Geobacter sulfurreducens, Shewanella putrefaciens, and Escherichia coli (Yang et al., 2019). To enhance the electron transfer between Shewanella oneidensis and the electrode, a novel strategy was adopted to fabricate a hybrid electroactive biofilm in which bacteria was inserted into the graphene-carbon-nanotube network and used as an anode (Zhao et al., 2015). The active bacterial communities in the denitrifying biocathode is complex and consists of both autotrophic and heterotrophic denitrifiers. A longterm operation of a denitrifying biocathode MFC revealed the most abundant phylotype in the cathode was Betaproteobacteria at the initial stage, which shifted to Gammaproteobacteria at the final stage (Chen et al., 2010). A study of the functional gene in the denitrification pathway revealed that the denitrifiers containing the nirS gene (nitrate reductase) were most abundant in the cathode biofilm (Vilar-Sanz et al., 2013). The effect of three types of inocula (denitrifying, denitratating, and ANAMMOX sludge) on the performance of denitrifying biocathode BES was investigated. Pseudomonas, Glycocaulis, and SM1A02 were found to be

3.2 Nitrogen removal and recovery

enriched in the three biofilms, and Pseudomonas stutzeri was presumed to be the most dominant bacteria, which utilized the electrons from the cathode for both ATP synthesis and denitrification. It was concluded that denitrifying sludge was most effective for wastewater treatment; however, ANAMMOX sludge prove to be optimal inoculum for combined wastewater treatment and energy recovery (Ding et al., 2018). It was also reported that exoeletrogens (Geobacter) and denitrifying bacteria (Thauera) can coexist in the anode of an air-cathode MFC and carry out both electricity generation and nitrate removal. The nitrite can be reduced by the bioelectrochemical nitrate reduction by the exoelectrogens or by the action of denitrifying bacteria (Huang et al., 2019). A recent study revealed that denitrifying bacteria could affect the electrogenesis by consuming organic carbon and electrons, thus altering the available substrate concentration. Geobactor was the most abundant anode-respiring bacteria, and Bacteroides, Dechloromonas, and Sterolibacterium were the denitrifying bacteria present in the anode (Su et al., 2018). An aerobic consortium of Thaueradominated denitrifiers was inoculated in the air-cathode MFC, and high electrochemical activity and pollutant removal were achieved. Thauera spp. have the ability to anaerobically degrade aromatic organic compounds, and it was assumed that it can transfer electrons to the electrode by membrane-bound proteins (Yang et al., 2019). Nitrate or nitrite can be directly converted to ammonium under anaerobic conditions via dissimilatory nitrate reduction to ammonium (DNRA). Geobacter lovleyi can perform DNRA utilizing nitrate as an electron acceptor. The positively charged ammonium nitrogen can be concentrated by electromigration for ammonia recovery (Li et al., 2019).

3.2.3 Ammonia recovery Due to the depletion of natural resources and high cost of nitrogen fixation, the recovery of nitrogen from the waste stream is considered as a more viable option than removal of it. Nitrogen is recovered in BES primarily as ammonia. PBS can concentrate nitrogen in algal biomass but its further utilization is challenging. Energy consumption of BES-based ammonia recovery was estimated at 4.5
7.5 kWh/kg N which is considerably lower than that of the Haber
Bosch process for ammonia synthesis (Qin et al., 2018). The recovered ammonia can be used for the production of fertilizers. The recovered ammonia could also be transformed into a solid form. It has been investigated in a three-compartment BES where ammonia was concentrated as ammonium bicarbonate salt in the middle compartment (Brewster et al., 2017). Solidifying ammonia would make it easier to transport and greatly expand the range of application. Ammonia can also be used as a substrate in ammonia fuel cells and generate electricity. This technique is currently under investigation and expected to be demonstrated in near future (Jain and He, 2018). This section focuses on ammonia recovery by migration and diffusion in BES. The recovery of ammonia by precipitation is discussed in the section 3.3.2 of this chapter. As discussed earlier, the ammonium ion can migrate through CEM from anode to cathode chamber via either current driven migration or diffusion. Electromigration of

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ammonium ion is a result of the electric field, while the diffusion of ammonia is induced by concentration gradient. The diffusion of ammonia (Cheng et al., 2013) and electricity-driven ammonium migration (Villano et al., 2013) is demonstrated in BESs. Due to high pH of the catholyte the migrated ammonium ion will be converted to aqueous ammonia and it could be recovered by NH3 stripping with a suitable gas stream. Thus, no additional alkalinity is required for increasing the catholyte pH (Arredondo et al., 2015). Ammonia that is vented from the cathode chamber can be passed through dilute sulfuric acid and recovered in the form of ammonium sulfate (NH3SO4) which can be reused as a fertilizer. Furthermore, the recovered ammonia can also be condensed and stored as liquid ammonia (Nancharaiah et al., 2016). A study demonstrated the formation of ammonium bicarbonate (NH4HCO3) by passing the ammonia recovered from the cathode of MEC through CO2, which was then utilized to draw solute for a forward osmosis (FO) system that extracted water from the MEC anolyte effluent. This state-of-the-art approach represents a synergistic connection between BES and FO techniques for multiple resource recovery at the same time (Qin and He, 2014). Ammonia loss from swine wastewater was studied in both aircathode MFC and dual-chamber MFC. Ammonia removal reached 60% after 5 days of operation in air-cathode MFC and 68% after 13 days of operation in dual-chamber MFC with ferricyanide as catholyte. The study concluded that ammonia loss is mainly due to its transport (by migration and diffusion) to the cathode and subsequent volatilization (Kim et al., 2008). A stripping/absorption-BES system demonstrated ammonia recovery from pig slurry in which ammonia migration enhanced when switched from MFC to MEC mode. The ammonia migration rate increased further when NaCl was used as catholyte under MEC mode. The pH of the catholyte increased while using NaCl and favored ammonia stripping/absorption (Sotres et al., 2015). It was reported that energy requirements for ammonia recovery in BES is comparatively lower than conventional ammonia stripping. BES is also advantageous because it does not require an alkali addition, and electricity can be generated (energy positive system) (Kuntke et al., 2012). To reduce the energy consumption for ammonia recovery in BES, passive separation of ammonia from cathode in absence of active aeration was investigated in an MEC. It was noted that active aeration could ensure higher ammonia recovery efficiency; however, the energy consumption for nitrogen recovery with passive separation (1.3 kWh/kg N) was significantly lower than that of (2.3 kWh/kg N) with active aeration (Qin et al., 2018). In a comprehensive numerical modeling for BES-based ammonia recovery, it was found that the chemical groups such as the NH3/NH1 4 group can buffer the elevated (or decreased) pH of the catholyte (or anolyte). Also, the transport of NH1 4 ions through CEM accounts for about 90% of the total current, thus confirming that the NH1 4 ions serve as efficient proton shuttles during MEC operations (Liu et al., 2016). A new BES (R2-BES) was developed to extract nutrients from wastewater. In the system, two internal cathode chambers with AEM and CEM was placed inside an anodic tank. Bioelectricity was generated from oxidation of organics in the anode, ammonium was migrated to the cathode driven by electric current, and phosphorus was removed via ion exchange with hydroxyl that was generated as a result of the cathode

3.2 Nitrogen removal and recovery

reduction. The R2-BES removed 52.4% of phosphate and 83.4% of ammonium nitrogen under an applied voltage of 0.8 V (Zhang et al., 2014a). A microbial electrolysis desalination cell (MEDC) was developed using multiple ion exchange membrane (IEM) pairs for the removal of organics and saline while concurrently recovering nutrients from municipal wastewater. Ammonium and phosphate ion concentrated in the product chamber through IEMs. The recovery efficiency of nitrate and phosphorus reached 66% and 63.7%, respectively (Li et al., 2020). The sewage and reject consist of high ammonium concentration but low organic load, so it is not suitable as an anolyte in BES. To recover the ammonia from reject water, BES was operated in the MEC mode where synthetic wastewater was fed to the anode and reject water was introduced to the cathode. For recovering ammonia, the hydrogen and ammonia gas evolved from the cathode chamber was passed through an acid adsorbent solution (2 M HCl). Ammonium recovery efficiency of 94% for synthetic reject water and 79% for real reject water was achieved (Wu and Modin, 2013). Roughly about 75% of nitrogen in domestic wastewater originates from urine. Urine is rich in key nutrients (N, P, and K) and typically contains 9 g NH1 4 2 N=L 1 (from 21 g Urea/L), 0.7 g PO32 2 P=L and 2 g K /L. Urea ((NH ) CO) is the main 2 2 4 nitrogen component in urine, which is hydrolyzed to ammonia (NH3) and carbamate (NH2COOH) by the enzyme urease. The carbamate is subsequently decomposed to ammonia and carbon dioxide. Therefore, every mole of urea present in urine releases two moles of ammonia (Arredondo et al., 2015; Nancharaiah et al., 2016). The viability of ammonia recovery in BES was examined in a dual-chamber MFC and it was concluded that diffusional flux enables ammonium ion transport until an equilibrium between the anode and cathode compartment is reached, whereas migrational flux facilitate the accumulation of ammonium against a concentration gradient (Kuntke et al., 2011). In the following study a dual-chamber MFC with gas diffusion cathode was used to recover ammonia from source-separated urine. The ammonia in the cathode chamber was volatilized due to high pH and removed by the gas stream used for aeration and subsequent adsorption into boric acid and dilute sulfuric acid. The MFC demonstrated a power output of 0.25 W/m2 and an ammonia recovery rate of 3.3 g N/m2 per day (Kuntke et al., 2012). Ammonia removal with simultaneous hydrogen production from urine was demonstrated in a MEC. It was concluded that optimization of organics conversion for electricity generation is necessary for the optimum ammonium migration, and also simultaneous removal of ammonia from the cathode is essential to promote ammonium diffusion over the membrane (Kuntke et al., 2014). The principle of MDC was applied to construct and MEC with IEM stacks to concentrate the ammonium and phosphate ion from diluted urine. The ammonium and phosphate (0.74 g/L NH1 4 2 N and 0.073 PO32 4 2 P) present in diluted urine were concentrated by a factor of 4.5 and 3, 32 respectively (3.34 g/L NH1 4 2 N and 0.21 g/L PO4 2 P) (Tice and Kim, 2014). A hybrid microbial electrolysis/electrodialysis cell was developed to recover nitrogen, phosphorous, and potassium from urine by bioelectroconcentration. The process (known as Ugold) requires B5 kWh/kg N to drive the electrochemical system (for aeration, recirculation, and pH control) (Ledezma et al., 2017). In a subsequent

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study, the hydrogen-producing cathode was substituted with an air breathing cathode with a suitable catalytic layer which enabled the system function at zero power input. The anode and cathode chambers were separated by a product chamber with a CEM between anode and the product chamber and AEM between the product chamber and the air breathing cathode. In the anode, the organics present in urine was converted to electron, protons, and CO2, and the effluent from the anode was fed to the CEM whereas bicarbonate and phosphate ions travel from cathode chamber through AEM, enabling the production of liquid fertilizer in the product chamber with no heavy metals (Freguia et al., 2019). It was reported that owing to the revenue generated from ammonium nitrogen recovery from urine and low energy input, the application of BES for urine treatment is economical (Arredondo et al., 2015). The performance of different BESs for removal/recovery of nitrogen are summarized in Table 3.1.

3.2.4 Challenges in nitrogen removal and recovery BES has several advantages over other existing technologies for nitrogen removal and recovery from waste. Further development in BES is necessary to overcome key challenges before its practical application. Although most of the current research focuses on the nitrogen removal, the recovery aspect can lead toward circular economy of nutrients. It is also important to prioritize between energy recovery and nitrogen recovery. Higher energy recovery will result in moderate electric current with slow migration of ammonia, whereas higher electric current will facilitate the transport of ammonia but will recover lower energy. So, a system should be designed for the desired output. It was pointed out that most of the research related to nitrogen recovery in BES provides power density (power output) data but lacks the total energy consumption data. It is important to consider the energy consumption by the system to decipher whether the system is energy efficient or not. This data will help in further development toward the commercialization of BES. In view of nitrogen removal in BES, the incomplete denitrification might generate nitrous oxide (N2O), which is a very powerful greenhouse gas (it has 298 times the global warming potential as CO2). Reducing the production of this gas in wastewater treatment is of great importance to minimize the impact on global warming and improve the denitrification efficiency. A sequential two-stage BES demonstrated complete denitrification and removal of accumulated N2O. However, more comprehensive study is required to adjust the operational conditions to improve bioelectrochemical denitrification and thereby reducing nitrous oxide emissions (Kelly and He, 2014; Cecconet et al., 2019). The removal and subsequent recovery of ammonia from the anolyte depends on variety of factors such as the ammonium concentration in the influent water, EAB community in the anode, type of IEM, electricity production, catholyte pH, removal of ammonia from cathode for ammonia diffusion, etc. Many of these factors are interconnected; for example, the catholyte pH is determined by the current flow, which in turn is dependent on the type of membrane and conductivity of the anolyte. The

Table 3.1 Removal and recovery of nitrogen in BES. Electrode material Sl. No. 1.

2.

3.

4.

BES configuration Dual-chamber MFC with external aerobic nitrification unit Dual-chamber MFC with SND in cathode Dual-chamber MFC with SND in cathode Three-stage RBC
MFC unit

5.

MFC with oxic and anoxic biocathode

6.

Dual cathode MFC

7.

Dual cathode MFC

8.

Dual cathode MFC

9.

MFC-MBR system

Wastewater/ electron donor

Anode

Cathode

C

Granular graphite rods

C

Granular graphite rods

C

Granular graphite rods

Granular graphite rods Granular graphite rods SS mesh

C

Carbon fiber brush

Graphite woven felt

Graphite felt

Graphite felt

CH3COONa and NH4
N as C and N source CH3COONa and NH4
N as C and N source

Carbon brush

Carbon brush

Carbon cloth

Synthetic municipal wastewater Glucose and NH4Cl as C and N source

Carbon brush

Carbon cloth and carbon brush Carbon cloth with Pt coating SS mesh and carbon felt

CH3COONa and NH4Cl as and N source CH3COONa and NH4Cl as and N source CH3COONa and NH4Cl as and N source CH3COONa and NH4Cl as and N source Synthetic wastewater

Granular graphite and graphite rod

Treatment efficiency

Nitrogen removal/ recovery

Power output

References

100% acetate removal 100% acetate removal 69.1% COD removal 93.9% COD removal 98.8% COD removal 93.7% COD removal 85%
99% COD removal

69.6% nitrogen removal

34.6 W/m3

Virdis et al. (2008)

94.1% nitrogen removal

16.6 W/m3

Virdis et al. (2010)

30.4% of the NH1 4 migration and 17.8% denitrification 63.8% TN removal

460 mW/m3

Sotres et al. (2016)




Sayess et al. (2013)

14 W/m3 for OMFC and 7.2 W/m3 for A-MFC


Xie et al. (2011)

66.7%
89.6% TN removal

2.9 W/m3

Zhang and He (2012b)




99.9% TN removal

294.9 mW/m2

Li et al. (2016a)

96.6%
97% COD removal

23.3 6 6.5% nitrate removal

1253 mW/m3

Zhang et al. (2014b)

97.4% NH1 4
N and 97.3% TN removal 44% TN removal

Zhang and He (2012a)

(Continued)

Table 3.1 Removal and recovery of nitrogen in BES. Continued Electrode material Sl. No.

BES configuration

Wastewater/ electron donor

Anode

Cathode

Treatment efficiency

Nitrogen removal/ recovery

Power output

References

99.6% COD removal 97% COD removal 89% COD removal 81.9% COD removal

87.6% TN removal

68 6 5 mW/m

Zhang et al. (2011)

98.7 6 1.8% ammonium removal 83% ammonium removal 95.5% TN removal

517 mW/m3

94.5% COD removal 71.9% COD removal


91.8% NH1 4 2 N and 90.6% TN removal




Jiang et al. (2019)

70.1% NO2 3 N and 63.2% total inorganic nitrogen removal 3.29 gN/d m2 ammonia recovery rate 57%
79% TN removal

2.67 mW/m2

Ge et al. (2020)

250 mW/m2

Kuntke et al. (2012)




Wu and Modin (2013) Zhang et al. (2014a)

10.

Sediment-type photo-MFC

Synthetic wastewater.

Carbon paper

11.

Algae cathode MFC

Landfill leachate

12.

Algae cathode MFC

13.

Algae biofilm MFC

Swine-farming wastewater Domestic wastewater

Carbon fiber brush Carbon cloth

14.

Photomicrobial nutrients recovery cell Constructed wetland-MFC

Synthetic wastewater

Carbon brush

Carbon paper with Pt load Carbon fiber brush Carbon cloth Carbon cloth with titanium wire Carbon cloth

Synthetic wastewater

Carbon fiber felt

Carbon fiber felt

16.

MFC with ammonia adsorption

Urine

Graphite felt

Pt coated Ti felt

17.

MEC with air stripping system

Reject water

Steel wire




18.

R2
BES

Synthetic wastewater

Carbon felt attached to graphite rod Carbon brush

Carbon cloth

79%
94.4% COD removal

15.

Carbon cloth with titanium wire

83.4% ammonia recovery

2

0.98 6 0.13 mW/ m2 62.93 mW/m2

Applied voltage was 0.8 V

Nguyen et al. (2017) Colombo et al. (2017) Yang et al. (2018)

19.

MEC with ammonia stripping

Pig slurry

Carbon felt mesh

SS mesh

20.

Tubular MEC

Synthetic wastewater

Carbon brush

Carbon cloth

21.

MFC with electrodialysis system Microbial electrolysis desalination cell

Urine

Graphite granules with graphite rod Carbon felt

Ti mesh

22.

Synthetic wastewater

SS mesh

46.8%
50.9% COD removal 89.1% COD removal


75.5% COD removal

49.9% ammonia migration and 94.3% nitrogen recovery




Sotres et al. (2015)

73%
90.1% ammonia recovery




Qin et al. (2018)

25
30 gN/d m2 ammonia recovery rate 66% nitrogen recovery




Freguia et al. (2019)




Li et al. (2020)

BES, Bioelectrochemical system; COD, chemical oxygen demand; MBR, membrane bioreactor; MFC, microbial fuel cell; RBC, rotating biological contractor; SND, simultaneous nitrification and denitrification; SS, stainless-steel; TN, total nitrogen.

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CHAPTER 3 Nutrient removal and recovery in bioelectrochemical

current flow is also affected by the high internal resistance of the BES, which could be optimized by modifying reactor configuration and improving electrode materials (Arredondo et al., 2015). Finally, pilot plant application is necessary to demonstrate the robustness and stability of the system in long-term operation and life cycle assessment (LCA) of the BES could assess the economic feasibility and environmental benefits of the system.

3.3 Phosphorus removal and recovery Phosphorus is a nonrenewable inorganic resource and important element in fertilizer for agricultural industry. Phosphorous is available in nature in the form of phosphate rocks. The phosphate rocks are unequally distributed in the Earth’s crust and mainly localized in a few countries. Due to its rapid consumption, depletion of the global minable phosphorous is expected by the end of 21st century (Gilbert, 2009). Phosphorous has gained attention as a valuable nutrient as well as a pollutant. The unutilized phosphorus can enter the natural water body as agricultural runoff and promote eutrophication. Wastewater generated by anthropogenic activities such as sewage, swine wastewater, and agro-industrial wastewater also contains phosphorous (Nancharaiah et al., 2016). Thus, phosphorous recovery from the waste streams and its subsequent utilization as a fertilizer is of great interest. A number of technologies have been explored in an attempt to recover phosphorous from nutrient-rich wastewater, such as adsorption (Westholm, 2006), ion exchange (Liberti et al., 1981), distillation (Udert and Wa¨chter, 2012), pyrolysis (Bridle and Pritchard, 2004), chemical precipitation (Wilfert et al., 2015), and biological uptake (Cai et al., 2013). However, these processes are often energy-intensive, and it is necessary to remove and recover phosphorus directly from wastewater with little (or no) energy input for the process to be sustainable. Phosphorous recovery in BES is primarily achieved by electroconcentration, struvite precipitation, and concentration in algal biomass (Fig. 3.4).

3.3.1 Issues related to biological phosphorus removal In wastewater treatment plants, phosphorous is removed by the action of a phosphate accumulating organism (PAO), and the process is termed as enhanced biological phosphorus removal (EBPR). EBPR removes phosphorous from the waste stream and accumulates it in a phosphorous-rich biomass, suitable for recovery and/or reuse. The phosphorous content in EBPR sludge (0.06
0.15 mg P/mg mixed liquor volatile suspended solids (MLVSS)) is significantly higher than that of traditional activated sludge (0.02 mg P/mg MLVSS). EBPR method commonly includes alternative anaerobic/anoxic/aerobic process (A2O) and volatile fatty acids (VFAs), such as acetate and propionate, which are added in the system for the enrichment of PAOs. EBPR depends on the capacity of PAOs to consume, convert, and accumulate excess phosphorus in the form of polyphosphate within the cells. Under anaerobic conditions, PAOs consume VFAs and store it as

3.3 Phosphorus removal and recovery

FIGURE 3.4 Schematic representation of different phosphorous removal mechanism in BES. (1) Reduction of iron phosphate compound in the cathode and release of phosphate; (2) removal of ammonium and phosphate by struvite precipitation; (3) phosphate uptake by algal biomass in a PBES; and (4) phosphorous recovery in MNRC. BES, Bioelectrochemical system; MNRC, microbial nutrient recovery cell; PBES, photobioelectrochemical system.

polyhydroxyalkanonate (PHA). The energy required by the PAOs to store PHA was provided by hydrolysis of polyphosphate to orthophosphate. The release of orthophosphate increases the soluble phosphorous concentration in the bulk solution. Under aerobic condition, the stored PHA is hydrolyzed to provide energy for glycogen synthesis and uptake of phosphorous for polyphosphate production in the cell. The phosphate uptake in an aerobic condition is higher than phosphate release in an anaerobic condition; excess accumulation of phosphorous in the biosolids separates it from the waste stream (Yuan et al., 2012). The surplus sludge contains toxic substances and pathogens, so its direct application in land is not recommendable. This EBPR sludge is stabilized in an anaerobic digester. Due to inadequate autotrophic metabolism of nutrients the digester effluent (digested sludge centrate) contains a high concentration of nitrogen and phosphorous, which is recovered as struvite. However, only around 50% of phosphorus can be recovered from influent wastewater because of the efficiency losses at EBPR and inefficient struvite crystallization under an imbalanced N/P ratio, leaving abundant amounts of unrecovered nitrogen in the centrate. Recycling the nutrient-rich centrate back to the aerobic biological treatment process leads to significant increase of nutrient loading rate and additional energy demand (Zou et al., 2017). Thus, development of an integrated process is required for sustainable nitrogen and phosphorous recovery from wastewater.

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3.3.2 Struvite precipitation The recovery of nutrients by chemical precipitation is regarded as one of the most promising techniques, owing to its high stability and efficiency. Calcium and magnesium compounds are generally used to induce precipitation. Magnesium ammonium phosphate hexahydrate (MgNH 4PO4.6H2O) or struvite is a white crystalline substance consisting of magnesium, ammonium, and phosphate in a molar ratio of 1:1:1 in slightly alkaline conditions. Under 32 21 supersaturation conditions, free NH1 4 , Mg , and PO4 ions and six molecules of water react together to produce struvite based on the following chemical reaction (Eq. 3.5) (Sciarria et al., 2019). 32 1 Mg21 1 NH1 4 1 PO4 1 6H2 O-MgNH4 PO4 6H2 Ok 1 2H

(3.5)

Similarly, hydroxyapatite (Ca5(OH)(PO4)3) also serves as an effective fertilizer supplement and can be precipitated by the reaction between calcium and phosphate described as Eq. 3.6 (Ye et al., 2019). 2 5Ca21 1 3PO32 4 1 3OH -Ca5 ðOHÞðPO4 Þ3 k

(3.6)

One of the advantages of struvite precipitation is the elimination of both phosphorous and nitrogen from the waste stream, resulting in an excellent fertilizer with low water solubility and gradual release in the soil. Additionally, the process is nonspontaneous and it could be utilized as a pretreatment or posttreatment depending on the wastewater characteristics. The major influencing factors for 32 struvite precipitation are pH, temperature, Mg21 :NH1 4 :PO4 molar ratio, and the 21 21 22 competing ions (e.g., Ca ; CO3 ). Generally, Mg ion is considered as a limiting ion as its concentration in real wastewater is comparatively lower than ammonium and phosphate ion concentration. Therefore, to achieve a desired molar ratio to induce struvite precipitation, external magnesium dosing (as MgCl2, MgSO4, and MgO) is required. Furthermore, the availability of the free ions depends on the H1 concentration, so pH can influence the saturation index of struvite and the kinetics of struvite precipitation. Usually, a pH more than eight is required for the reaction to happen. Commonly, pH of the real wastewater is lower than eight, and pH adjustment is necessary to promote struvite precipitation. The conventional method of pH adjustment includes the addition of base (such as NaOH, Ca(OH)2, Mg(OH)2) and CO2 stripping by aeration (Sciarria et al., 2019; Lin et al., 2018; Kim et al., 2018). Nutrients can be recovered in BES as struvite (such as MFC and MEC) by taking advantage of the high pH of catholyte due to current flow and hydroxyl ion formation. Struvite can be recovered near or on the cathode without the addition of alkaline chemicals. It is noteworthy that an excessively high pH will adversely affect struvite crystallization due to the formation of magnesium hydroxide (Mg(OH)2) and magnesium phosphate (Mg3(PO4)2) instead of struvite (Lin et al., 2018). Hence, the pH value of catholyte should be closely monitored for struvite precipitation.

3.3 Phosphorus removal and recovery

3.3.3 Phosphorus removal and recovery in bioelectrochemical system BES may be used to solubilize phosphorus from iron phosphate minerals (Fe-P), as well as to precipitate soluble phosphorus in the form of struvite (Sun et al., 2018). Phosphate exaction from digested sewage sludge (containing FePO4) and its subsequent recovery as struvite was demonstrated in a dual-chamber MFC. The electrons generated in the anode were utilized to reduce FePO4 in the sludge, and phosphate was released in the catholyte. Stoichiometric amounts of magnesium and ammonium was added in the cathode as phosphate was removed by struvite precipitation (Fischer et al., 2011). Remobilization of phosphate from digested sewage sludge containing Fe-P was also demonstrated in a three-literthree-chambered MFC. To accelerate phosphate solubilization from Fe-P, the operation of the system switched from MFC to MEC mode. Application of electric current increased the pH of the catholyte to 12.6, which accelerated phosphate remobilization rate and 67% phosphate was recovered within 26 hours which was significantly faster than MFC conditions. The remobilized phosphate was recovered as struvite by precipitation using an equivalent amount of MgCl2 and NH4OH (Happe et al., 2016). Simultaneous hydrogen production and struvite precipitation was demonstrated in a single-chamber MEC in which achieved soluble phosphate removal of 40% with struvite crystallization rate of 0.3
0.9 g/m2 per hour (Cusick and Logan, 2012). Phosphorous recovery from swine wastewater has been studied in an air-cathode MFC, in which 70%
82% of the influent phosphorus was removed and struvite precipitation observed on the cathode surface exposed to liquid phase. Even though the effluent pH was around 8.03, the oxygen reduction in cathode presumed to increase the localized pH on the cathode surface, facilitating struvite precipitation (Ichihashi and Hirooka, 2012). Struvite precipitation was studied in a single-chamber air-cathode MFC by adding magnesium, ammonia, and phosphate in synthetic wastewater. It was observed that in the absence of ammonia, phosphate removal efficiency declined. Magnesium and phosphate were removed by cattiite (Mg3(PO4)2) precipitation. Also, the addition of NaOH facilitated struvite precipitation by elevating the pH to 8.62. Moreover, struvite precipitation on the cathode surface hindered the mass transfer of ions and oxygen, resulting in deterioration in MFC performance. After removal of the precipitate, the cathode performance was restored to the initial level (Hirooka and Ichihashi, 2013). Simultaneous nutrient, energy, and water recovery was achieved in a novel MEC-FO system. The hybrid system recovered B54% of water, B99% of net ammonium nitrogen (by extended N2 stripping), and B80% of phosphorous from digestion centrate (Zou et al., 2017). Struvite is crystallized on cathode surfaces; hence cathode configuration plays a significant role in the efficient struvite precipitation in BES. A comparison of cathode materials for enhanced phosphorous recovery from dewatering centrate as struvite was demonstrated in a MEC. The study recommends foil type cathodes over mesh (stainlesssteel) cathodes to prevent possible losses of small struvite crystals. It was also noted that the unavailability of readily biodegradable organics in the dewatering

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centrate resulted in low current output, which was insufficient to raise the pH near the cathode for complete recovery of phosphorous as struvite. The addition of acetate in the anode recovered up to 96% of phosphorous using stainless steel foil as cathode (Yuan and Kim, 2017). Research has also found that a low-cost, stainless-steel mesh cathode can give comparable output as a platinum electrode in MFC and MEC while treating digestate and simultaneously producing electricity and H2. It has also demonstrated that seawater bitterns, a low-cost natural source of Mg21 ions obtained as a byproduct from salt processing, can promote struvite precipitation replacing the usage of pure salt (MgCl2.6H2O) (Sciarria et al., 2019). DO concentration in the catholyte also affects the performance of the BES. A reduction in DO in the cathode chamber may restrict the electron flow affecting the ammonia migration from anode to cathode as well as reduce the catholyte pH resulting inefficient struvite precipitation (Tao et al., 2014). An extensive study was done in an MFC comparing different design operating parameters (operational mode, aeration condition in the cathode, and type of membrane) to achieve optimal nutrient removal and recovery from municipal wastewater. Sequential anodic and cathodic treatment of the wastewater facilitated phosphate removal by microbial absorption and recovery by chemical precipitation, while ammonium was accumulated by electromigration and retrieved as precipitates (Ye et al., 2019). A two-step process involving MEC and chemical precipitation was investigated for the removal of nutrients from digested sludge centrate. Ammonia was recovered in the cathode of a dual-chamber MEC with simultaneous hydrogen production and phosphorous was recovered as precipitate by the addition of magnesium chloride and calcium chloride in the anode effluent. The low organic content of centrate hindered the current generation and ammonia migration. As a potential source of biodegradable organics, the addition of primary sludge fermentation liquor significantly increased current density and subsequent ammonia recovery in MEC. For phosphorous recovery as calcium phosphate, no pH adjustment was required (anolyte pH 6.8), whereas struvite precipitation required anolyte pH adjustment to 8.6 with sodium hydroxide. Further compressive evaluation of different options for phosphorous precipitation is required to maximize nutrient recovery from digested sludge centrate (Barua et al., 2019). As discussed earlier, PBES can remove nitrogen and phosphorous from wastewater and concentrate it in algal biomass. A sediment-type PBES was investigated for nutrient removal from synthetic wastewater based on the synergistic interaction between microalgae (Chlorella vulgaris) and EAB. Almost 70% phosphorous was removed with 99.6% organics and 87.6% nitrogen removal (Zhang et al., 2011). Algal bioreactor is also used to treat the effluent of MFC to remove the remining nitrogen and phosphorous by microalgae. Inclusion of external photobioreactor improved the phosphorous removal from 58% to 92% (Jiang et al., 2012). Nutrient removal with simultaneous electricity generation from wastewater was investigated by installing MFC inside an algal bioreactor. The system achieved 98% of ammonium nitrogen, 82% of phosphate, and 92% of COD removal and produced a maximum power density of 2.2 W/m3 (Xiao et al., 2012). Further utilization of algal

3.3 Phosphorus removal and recovery

biomass for phosphorous recovery requires thorough investigation. For simultaneous removal of COD and nutrients, a technology termed as microbial nutrient recovery cell (MNRC) was developed. In the MNRC, a recovery chamber separates the anode and cathode chambers; CEM is placed at the anode side and AEM is placed at the cathode side of the recovery chamber. Wastewater is recirculated between the anode and cathode, and due to current flow, the ammonium from anode and phosphate from cathode was migrated into the recovery chamber. More than 82% COD, 96% ammonium nitrogen, and 64% phosphate was removed from the wastewater. The concentrations of ammonium and phosphate was 1.5 and 2.2 times more than its initial concentration in the influent wastewater (Chen et al., 2015). Nutrient removal and purification of urine was achieved in a BES called U-Power, which provided a maximum power density of 21.3 W/m3 by degrading fresh human urine. Urea present in the urine was hydrolyzed by urease produced by anodic microorganism. Electrical potential generated by U-Power facilitated the ammonium and phosphate migration to the recovery chamber. On average, 93.8% COD, 73.1% TN, and 86.2% total phosphorous removal from urine was achieved, while 1.2 g/L TN and 0.1 g/L total phosphorous was recovered in the highly concentrated recovery solution. U-Power constitutes a promising method to pave the way for sustainable water-energy-nutrient nexus (Gao et al., 2018). The phosphorous removal performance of different BESs is summarized in Table 3.2.

3.3.4 Challenges in phosphorus removal and recovery The recovery of phosphorous through BESs is still in infancy stage of development. Phosphorous recovery in BES is not studied as extensively as nitrogen recovery as phosphorous is mainly recovered through precipitation, and it is not related to the electricity generation mechanism. Future investigation is necessary to understand the effect of phosphorous removal on current production in BES. The recovery of phosphorous in single-chamber BES can be hampered by the pH of the system. As struvite precipitation requires alkaline condition, the acidic pH of the anolyte can prevent the formation of struvite. The struvite precipitation in a small amount can only occur on the surface of cathode where the increase of localized pH can facilitate favorable conditions for struvite precipitation. For efficient removal of phosphorous dual-chamber configuration or MNRC with a recovery chamber has proved more efficient. It was also observed that the deposition of struvite on the cathode surface could affect the performance of the BES by imposing mass transfer limitation. Frequent collection and replacement/regeneration of the cathode can be a significant challenge in real-world applications. Moreover, most of the research focuses on phosphorous recovery as struvite, which requires the presence of a stoichiometric amount of ammonia and magnesium. The unavailability of these key constituents in the wastewater would result in inefficient phosphorous removal. On the other hand, addition of these chemicals can increase the cost of treatment. The precipitation of phosphorus in some other forms, such as calcium phosphate, needs extensive investigation as it can

71

Table 3.2 Removal and recovery of phosphorous in BES. Electrode material Sl. No.

BES configuration

Wastewater

Anode

Cathode

Recovery mechanism

Phosphorous removal/recovery

Power output

1.

Single chamber MEC

Synthetic wastewater

Graphite fiber brush

SS mesh

Struvite crystallization




2.

Air-cathode MFC

Swine wastewater

Carbon felt

Struvite precipitation

3.

Single chamber MFC

Artificial wastewater

Carbon felt disc

Dual-chamber tubular MFC

Synthetic wastewater

Carbon paper

5.

Three chambered MFC-MEC system Single-chamber MEC Air-cathode MFC

Sewage sludge

Reticulated vitreous carbon

Graphite plate

Remobilization of phosphate and precipitation as struvite

19%
55% P removal as precipitates 90% total phosphorous (TP) removal 67% phosphate remobilized from FeP




4.

carbon paper with Pt coating Carbon paper with Pt coating Carbon cloth

40.3% Phosphate removal and precipitation 70%
82% P removal

Graphite fiber brush Graphite fiber brush

SS mesh/foil

Struvite precipitation

SS mesh

Graphite felt Carbon paper

carbon fiber brush with Ti coating Carbon paper

6.

8.

Dual-chamber MFC

Dewatering centrate Digestate from anaerobic digester Municipal wastewater

9.

Sediment-type photo-MFC

Synthetic wastewater

7.

Struvite and cattiite precipitation Chemical precipitation

2.3 W/m2

530 mW/m2

References Cusick and Logan (2012) Ichihashi and Hirooka (2012) Hirooka and Ichihashi (2013) Tao et al. (2014)




Happe et al. (2016)

92% P removal




Struvite crystallization

35.8% P removal in MFC and 83.1% P precipitation

14.2 W/m3

Yuan and Kim (2017) Sciarria et al. (2019)

Struvite precipitation

94.9% of PO32 4
P removed/recovered




Ye et al. (2019)

Algal uptake

69.8% P removal

68 mW/m2

Zhang et al. (2011)

10.

11.

12.

13.

MFC and photobioreactor system Single-chamber tubular PBES

Domestic wastewater

Carbon fiber brush

Carbon fiber brush

Algal uptake

58% TP removal

20.3 W/m3

Jiang et al. (2012)

Synthetic wastewater

Carbon brush

Algal uptake

82% of phosphate removal

2.2 W/m3

Xiao et al. (2012)

Microbial nutrient recovery cell Microbial nutrient recovery cell

Synthetic wastewater

Titanium mesh

carbon cloth with Pt/C catalyst Carbon cloth

64% phosphate removal




Chen et al. (2015)

Urine

Granular activated carbon

Electromigration and collection in the recovery chamber Electricity driven migration

86% TP removal

21.3 W/m3

Gao et al. (2018)

SS mesh

MEC, Microbial electrolysis cell; MFC, microbial fuel cell; PBES, photo-bioelectrochemical system; SS, stainless-steel.

74

CHAPTER 3 Nutrient removal and recovery in bioelectrochemical

also be utilized as fertilizer. The precipitation of calcium phosphate occurs in pH 6.8, so it can be precipitated from anode effluent without any pH adjustments. Another important aspect to consider is nutrient-rich wastewater (e.g., digester centrate) lacks a suitable quantity of biodegradable organics, which is required by the anode-respiring bacteria for current generation. It might hinder the migration of ammonia from anode to cathode and reduction reaction in cathode to elevate the pH level, resulting in inefficient struvite precipitation and inadequate phosphorous removal. The addition of carbon sources (such as acetate) in the wastewater may incur extra cost and increase complexities of wastewater treatment by additional COD load. The blending of nutrient-rich low organic content wastewater with high strength wastewater and its subsequent treatment could solve this problem. For realizing the process at an industrial scale, studies should also focus on the long-term performance of BES in terms of removal and recovery efficiency, energy recovery, stability, and financial viability of the system.

3.4 Conclusion and future perspectives Simultaneous nutrient removal and electricity generation in BESs make them more advantageous over other technologies. Also, nutrient recovery from different waste streams can address rising global concerns over the depletion of natural mineral resources. However, it is still in the developmental stage and constrained by various operational and economic challenges, and it requires further effort since the efficiency, robustness, long-term stability, and expenses remain uncertain. In BES, the current generation plays an important role in both nitrogen removal/recovery and coulombic efficiency. Owing to the limitations of the electricity generation capacity of BES, nitrogen recovery should be prioritized over energy recovery. The high current flow will facilitate ammonia migration and pH elevation, but energy recovery will be limited. Also, with limited energy recovery, effective wastewater treatment is not achievable. A BES can be designed comprising of multiple modules with different functions where nitrogen recovery and energy recovery can occur in separate reactors. The energy generation in BES largely depends on EAB community and the stability of the biofilm over anode surface. The effective electron recovery can enhance both conversion efficiency and product recovery. So, anode configuration and operational parameters need to be monitored to ensure the potency of the electroactive biofilm. Wastewater/waste characteristics also plays a significant role for efficient nutrient recovery. For example, wastewater with low organic and high nutrient content needs to be supplemented with additional electron donor to enhance the recovery efficiency. On the other hand, when both the carbon source (electron donor) and nitrate (electron acceptor) are present in the wastewater, denitrification can occur without the BES system. In this scenario, it is beneficial to convert nitrate into ammonia by DNRA bacteria, and subsequently, it can be recovered via electromigration in the cathode of the BES. In this regard, it is worth mentioning that the ammonia should be

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Role of bioelectrochemical systems for bioremediation of wastewaters and bioenergy production

4

Muhammad Faisal Siddiqui1, Zahid Ullah1, Lakhveer Singh2, Farhana Maqbool1, Sadia Qayyum1, Ihsan Ullah3, Ziaur Rahman4 and Fazal Adnan5 1

Department of Microbiology, Hazara University, Mansehra, Pakistan Department of Environmental Science, SRM University-AP, Amaravati, India 3 Department of Biological Sciences, Faculty of Science, King Abdulaziz University, Jeddah, Saudi Arabia 4 Department of Microbiology, Abdul Wali Khan University Mardan, Khyber Pakhtunkhwa, Pakistan 5 Atta ur Rahman School of Applied Biosciences, National University of Sciences & Technology, Pakistan 2

4.1 Introduction The need for alternative energy sources is increasing because of the restricted supply and pollutants generated from different industries. The worldwide current demand of energy stands at around 13 terawatts (TW), and it is estimated that it can be reach around 23 TW in 2050 (Chae et al., 2009; Singh et al., 2013a). Furthermore, extensive use of resources in agricultural, municipal, and industrial activities will evaluate the environmental degradation process by creating threats to the environment, that is, global warming and associated consequences. Advancing renewable sources of energy is becoming important for an environmentally friendly ecosystem, particularly reduction depending on the imports of energy and expanding sources of energy generation (Singh et al., 2013a,b,c; Resch et al., 2008). Over the last decade, bioelectrochemical systems (BESs) have boomed considerably because of their importance by means of an evolving, efficient technology, not only for energy production but also for remediation of wastewaters (Mohan et al., 2010). In addition, BESs often give excellent opportunities for the safe and effective use of microorganisms to process high-value chemicals and fuels (ElMekawy et al., 2015). In practical terms, BES can be said to be traditional electrochemical systems by using microbes as catalysts (Baeza et al., 2017). Energy generated by a few bacteria was already known long ago (Potter, 1911). This trend did not start to attract significant attention of scientists and Delivering Low-Carbon Biofuels with Bioproduct Recovery. DOI: https://doi.org/10.1016/B978-0-12-821841-9.00010-4 © 2021 Elsevier Inc. All rights reserved.

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engineers until the beginning of the present century. The gradual development in BESs from last 15 years has enabled the pilot-scale production transformation from lab-scale experiments (Baeza et al., 2017; Cotterill et al., 2017). And it seems that commercial growth is closer. The initial research activities centered on finding the opportunities offered by these electrochemical systems for energy recovery as well as for remediation of waste waters (Rozendal et al., 2006; Liu et al., 2004). Because the BES operates with a fully multidisciplinary manner, a significant amount of research has been carried out globally in different fields including biotechnology, electrochemistry, microbiology, environmental science, and bioelectrochemistry (Patil et al., 2015). The electrons traveling into an external circuit are considered electrical current. The mechanism is referred to as the microbial-based electrochemical process (Harnisch and Schro¨der, 2010). It is known that microbial cells produce energy in anode chamber from biodegradation of organic waste materials. For production of bioenergy, low redox potential is needed for oxidation of organic wastes, however, for oxygen reduction reaction, high redox potential is needed (Singh et al., 2015). Microbial electrosynthesis (MES) systems are usually those systems in which organic molecules are cathodically reduced (Rabaey and Rozendal, 2010). Moreover, microbial cells are used with combination with other technologies; for example, desalination cells or microbial fuel cells using plants in which roots of plants are employed for electricity generation (Luo et al., 2012; Chiranjeevi et al., 2012; ElMekawy et al., 2014b; Mohanakrishna et al., 2010; Siddiqui et al., 2019). Also, bioelectrochemical systems are applied as efficient bioreactors for the remediation of toxic compounds; this technique is known as microbial bioelectrochemical treatment of wastewater (Mohanakrishna et al., 2010).

4.2 Principle of bioelectrochemical systems The BES theory is focused on electrochemical transformation processes, which through catalytic action of microorganisms transform the chemical energy contained in biodegradable organic matter. BES is comprised of cathodic and anodic cells, which are separated by membranes called exchange membranes, which are used for the oxidation reaction and for the reduction reactions in the cell (Rabaey and Rozendal, 2010). The bacterial cells that are activated perform oxidation reaction in the anodic chamber of organics and produce electron donors. These electron donors are arrested by energy/electricity via transferring electrons to cathode using external circuit (Wang and Ren, 2013). Due to its tremendous advantages and uses of these systems, for example, in remediation of different toxic wastewaters and for production of bioelectricity and many valuable products productions, and due to these incredible applications, bioelectrochemical systems have gained massive attention from many researchers over the last two decades.

4.3 Kinds of bioelectrochemical systems

4.3 Kinds of bioelectrochemical systems There are various types of BESs, and these systems depend on the need of the users. These systems include; bioelectrochemical treatment systems, microbial desalination systems based on desalination, electrogenesis systems, etc.

4.3.1 Microbial fuel cells Microbial fuel cells (MFCs) use a solid electrode as an electron-acceptor to arrest current, which is produced due to oxidation of organic waste material with the help of microbial cells (Rabaey et al., 2011). It is known that microorganisms, especially bacteria, help in the oxidation of organic compounds, and then they produce electrons, which are transferred to the cathode chamber and hence captured by cathode. Electron transfer occur with the help of external circuit. Potter (1911) had shown the first of these devices attaining a high voltage of 0.3-0.5 V. It was found that the microbial biofilm is attached on the anode in the anode chamber, and this biofilm act as electrocatalytic parts to produce electricity. The production of electricity using microbes is dependent on the exoelectrogen efficacy of microbial cells attached on the anode part. The anode helps in the transferring of electrons generated from the oxidation of organic substrates. Diverse microbial species allow the oxidation of different substrates in the anode chamber. It was observed that Shewanella was helping in the oxidation of substrates. Furthermore, Geobacteraceae’s were also investigated and they were also producing high power as reporting in different studies (Kiely et al., 2011; Loman et al., 2012). It was find out that the high current density was 390 Am2 in MFCs in which a mixture of microbes were utilized (Chen et al., 2012). It was observed that the chemistry of the electrolyte, nature of anode, and chemistry of organic substrate affects the whole process, and it also affects the movement of electrons from anodic chamber (Fan et al., 2008; Pant et al., 2010). It was also observed that improved efficiency of microbial systems was found when high electrolyte conductivity was observed (Sharma et al., 2015). However, it was also noticed up to 85 Am2 of current densities was obtained from salty water (higher concentration of salts than sea water) when halophilic bacteria was used in BESs (Rousseau et al., 2013). MFC output often depends on the substrate used for the electrode and its composition. For a rough base, graphite electrodes have demonstrated higher power (Buisman et al., 2008). Similarly, it was found that the anode surface region of carbon nanotubes and graphite brushes have increased considerably in present density (Chen et al., 2011; Feng et al., 2010; Mohanakrishna et al., 2012). A higher electrode surface area offers a greater large surface for attachment of bacterial cells and exhibiting significant increases in the current (Dumas et al., 2008; Pocaznoi et al., 2012). It is observed that anodes made of carbon (porous) have high densities of current compared to metals (Zhang et al., 2014). A change to 3D

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electrodes with engineered structures was associated with a substantial improvement in the efficiency of bioelectrodes (Sharma et al., 2014). Modification of the macrostructure demonstrated a significant increase in existing current (Chen et al., 2012). Catalysts, for starters higher power densities recorded when Magnesium oxide coated cathodes were used in Microbial fuel cells. Materials used in the building, design and configuration of MFCs affect the internal resistance, which in effect determines the overall device output.

4.3.2 Microbial electrolysis cells for energy Microbial electrolysis cells (MECs) use bacterial properties to transform biochemical energy into electrical, thus permitting electrolysis of water (Verstraete and Rabaey, 2006). Synthesis of hydrogen occurs at cathode due to the movement of electrons from cathode anode because of external pressure added to the BES electrical circuit (Jeremiasse et al., 2010). Mostly, these cathodes work best under the anaerobic environment. Moreover, with hydrogen demand, may also facilitate the demand of methane until CO2 and methanogens are accessible. Many of the approaches to reduce the development of methane include aerating the chamber of having cathode, also it can be achieved by decreasing the pH and reducing the retention time (Logan et al., 2008).

4.3.3 Microbial electrosynthesis for energy production MES uses the power generated by oxidation in an anodic chamber. This system is a new kind of system. Attached cathodic biofilms minimize the volume of electrons necessary to manufacture added-value goods (Rabaey and Rozendal, 2010; Marshall et al., 2012; Mohanakrishna et al., 2015). The mechanism of bioelectrosynthesis may be extremely precise when biological catalysts help in the redox reaction. In this system, the electron acceptor helps in the cycle together with the redox meditators. It is worth noting that biological cathodes in this system are the necessary component. These microbes also help in the production of important products such as butyrate or ethanol, etc. (Bajracharya et al., 2015).

4.3.4 Enzymatic fuel cells for energy production These systems are also known as enzymatic fuel cells (EFCs). In this system the core component is the use of specific enzymes. Two different approaches are used in which in one approach enzymes must be attached on the surface of electrode and in another approach, it must be in suspension in the electrode chamber. These enzymes facilitate oxidation and other important reactions in the cell (Davis and Yarbrough, 1962), demonstrating the usage of glucose oxidase in bacteria and enzymes in biofuel cells. Due to thermostability and strong selectivity of glucose oxidase, glucose oxidase was commonly used in pacemakers, warning

4.3 Kinds of bioelectrochemical systems

lamps, miniature (Ieropoulos et al., 2008). Several enzymes, depending on their unique redox role, are used on anode and cathodes. It is observed that when enzymes are used as catalysts in these cells, they partially oxidize the fuel; due to this, heat is generated because of the other reactions (Minteer et al., 2007; Palmore and Whitesides, 1994). Furthermore, the life span of these enzymes is very short, so it is increased by capsulating enzymes in polymers which have a function of a hydrophobic niche to avoid enzymatic degradation (Minteer et al., 2007). Different enzymes are tested, such as glucose dehydrogenases (Zafar et al., 2012) and alcohol dehydrogenases (Gellett et al., 2010; Yakushi and Matsushita, 2010). Various redox mediators are used by alcohol dehydrogenases (Yakushi and Matsushita, 2010). However, the cost to EFCs is still large because of the usage of distilled enzymes with catalytic ability.

4.3.5 Microbial solar cells for energy production In these systems plants or photosynthesis microbes are used for absorption of energy from light or solar energy, which is then used for electrode-driven reactions by electroactive bacteria. These reactions involve electrical current generation, or substances such as carbon, gasoline, ethanol, hydrogen peroxide, etc. Photosynthesis contributes to the production of compounds, which are then used in the anode chamber. Electroactive microbes oxidize them to create electrons (ElMekawy et al., 2014a). Chlorophyta and other microorganisms are also attached on the anode as biofilms for this purpose (Strik et al., 2011). These systems are useful for the power generation and treatment of wastewaters (Wang et al., 2012). Upon care in an anaerobic digester, photobioreactor algal biomass may be used as a fuel cell feed (De Schamphelaire and Verstraete, 2009).

4.3.6 Plant microbial fuel cells for energy production This is an interesting system in which plant roots are integrated into the anode and in which deposits are produced by the root system; these compounds are then used as substrates and also acted on by different microbes (Timurkutluk et al., 2011). Therefore, these systems arrest radiation from sunlight, and this system cleanly and efficiently turns it into renewable energy. It was noted that the root rhizodeposits comprise primarily of different organics like amino acids and sugars (Timurkutluk et al., 2011). It is important to select a proper plant for these systems; different plants were used, such as Spartina angilica (Helder et al., 2012) and reed plant (Timurkutluk et al., 2011). It was further observed that Pennisetum setaceum was efficient in producing energy and was good for power generation up to now, with a maximum production of 163 mW m2 (Chiranjeevi et al., 2012). Plant microbial fuel cells (PMFCs) with optimized biocathodic oxygen reduction achieved a cumulative long-term power production of 240 mW m2 (2 weeks) (Wetser et al., 2015). PMFC is regarded as a successful and innovative energy harvesting device along with MSCs providing phototrophic biofilms.

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4.3.7 Microbial desalination cells for energy production These plants utilized the potential specially electrical which is formed in the system and utilized MFC technology to desalinize (Luo et al., 2012). In these systems there is an extra chamber with the middle chamber for water desalination. This portion is apportioned toward the anode via a member. Bacteria helps the anode oxidize the organics, and due to this, protons and electrons are produced. Then, electrons are transferred to the cathode chamber to cathode; however, interestingly, the anions generated are transferred to anode. Due to this, it preserves the equilibrium of the charge and thereby desalinate the middle chamber solution (Luo et al., 2012; Luo et al., 2011). Ions from saltwater in microbial desalination cells (MDC) improves the anolyte and catholyte conductivity. Because of higher conductivity and mass transmission, electrical power output in MDC was therefore increased. Moreover, there is limitation of this system that due to higher salinity it can affect the process of cathodic and anodic bacteria (Luo et al., 2012).

4.4 Role of bioelectrochemical systems in remediation of pollutants 4.4.1 Remediation of organic xenobiotics 4.4.1.1 Azo dyes remediation Wastewater from textile dying industries produces a massive amount of pollutants in the water bodies (Siddiqui et al., 2020; Siddiqui et al., 2017; Siddiqui et al., 2012), and it causes serious problems for the living organisms and severely deteriorates the environment. Among the dyes, azo dyes, which are organic dyes (organic compounds), have been studied in bioelectrochemical systems for remediation. Due to reduction, azo dyes are degraded into different components like amines and other compounds. So it was found that azo compounds act as electron acceptors and reduced in the system (Sun et al., 2009). It was observed that brilliant red dye was degraded when a mixture of bacteria was used from the sludge under aerobic and anerobic environment. Since then, other experiments have shown that electroactive microorganisms can degrade them to their constituent amines using a variety of different azo dyes. In a basic analysis of history, above all among these (Fernando et al., 2012) it was demonstrated for the first time that Shewanella oneidensis, an electroactive microorganism, was found to biodegradation acid orange. Furthermore, it was also found that mixture of azo dyes can be remediated and degraded in a two-chamber MFC by mixture of bacteria (Fernando et al., 2013). Bioremediation and biodegradation process of these dyes, especially azo dye, is widely agreed through a reduction of the azo chromophore (Fernando et al., 2012). It was first demonstrated that azo dye removal followed higher kinetic speeds in the presence of electroactive microorganisms (Fernando et al., 2012).

4.4 Role of bioelectrochemical systems in remediation of pollutants

4.4.1.2 Nitrobenzene compounds remediation Nitrophenol compounds were degraded and converted into the corresponding amine by introducing electrons and protons (Feng et al., 2011) Abiotic-cathodic degradation of 2-nitrophenol has been shown (Feng et al., 2011) and removal of parafluoronitrobenzene by a biocathode was possible (Feng et al., 2011). It was also exhibited that Fe (II)/Fe (III) introduction in the system can increase the reduction process of nitrophenol in the cathodic abiotic chamber. This biotransformation of nitrophenol in different components like amines can be just partial transformation, but the subsequent conversion of the produced products is more prone to biodegrade; it may be due to these compounds being toxic for bacterial activity (Minteer et al., 2007).

4.4.1.3 Chloronitrobenzene remediation These compounds, which are called chloronitrobenzene, are considered as poisonous and resistant to degradation. They are used in different sectors, ranging from coloring, weapons, and pharmaceuticals (Jiang et al., 2016). However, a recent analysis using BESs used up-flow reactor methods and found that these compounds can be converted effectively into comparatively more biodegradable end items. Of special interest are chlorinated aromatics that are immune to biotransformation. Consequently, a more important consequence of this research was that 2,4-dintrochlorobenzene was also dechlorinated. A microbial population sample review of the aforementioned study revealed that upflow anaerobic sludge blanket (UASB) comprising BES, in which most of the microbes present, were Desulfovibrio, Archobacter, and specie acetobacter were (Lu et al., 2012; Logan, 2009; Freguia et al., 2010).

4.4.1.4 Remediation of polychlorobiphenyl pollutants Many studies (Aulenta et al., 2007; Aulenta et al., 2011; Chun et al., 2013) revealed electrochemically active microorganisms in BES can effectively bioremediate polychlorinated biphenyls (PCBs) and halocarbon compounds. In all cases (Jiang et al., 2016; Aulenta et al., 2007; Aulenta et al., 2011), BESs working with a regulated cathode capacity, varying from 2250 to 2750 mV, have been shown to be able to rapidly perform microbially mediated reductive dechlorination of trichloroethene (TCE)—a halocarbon compound. In the same two experiments it was shown that when the cathode was roughly 2250 mV, it was observed that suppression in methanogenesis activity was noted. A similar study was also found that using a BES system, in which common halocarbon cisdichloroethene was reduced by bioremediation (Aulenta et al., 2013). The PCBs are considered to be extremely hazardous and cancerous contaminants in the atmosphere (Kannan et al., 1988). In a recent study conducted by (Chun et al., 2013), it was found that PCB degradation was demonstrated in Benthic BES structures. The most relevant element of the research done by (Chun et al., 2013) has been the in-situ PCB degradation accomplished under nonlaboratory conditions with an applied voltage (1.5 3 V). Usage of higher voltages supported initial

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dechlorination, accompanied by chlorobenzoate production and subsequent intake— indicating degradation of these pollutants. In some previous experiments, utilizing soil-related BES, they have also been successfully depleted and detoxified. Chlorinated organics were considered to be extremely resistant to degradation, and these compounds have a long half-life in the atmosphere (2.6 22.9 years in soil) (Cao et al., 2015). In the study conducted (Cao et al., 2015), 71% remediation was achieved in 56 days in BES, and it was observed that reduction was reduced from 40 to 200 mg/kg soil. Their rapid degradation in BES (as contrasted with traditional systems) is very important. The current study exhibited the exciting ability of these systems for biotransformation of recalcitrant environmental toxins effectively and rapidly. They also showed that, contrary to conventionally used physicochemical approaches of environmental remediation, which can be accomplished in a fairly environment friendly way via the use of BES-dependent structures.

4.4.1.5 Polyaromatic hydrocarbons and related compounds remediation In recent studies, petroleum hydrocarbons were demanding pollutants to be degraded due to its structure and complex nature of these pollutants. Petrochemical compounds are known to be resistant to biodegradation, particularly polycyclic aromatic hydrocarbons (PAHs). The usage of hydrocarbon emissions as an electron donor in MFC systems is an interesting opportunity, where corresponding hydrocarbon-pollutant power generation and oxidation can be accomplished. It was proven first by Morris and Jin (2007) and Morris et al. (2009). The anaerobic biodegradation by MFCs (mixtures of aliphatic hydrocarbons varying from C-8 to C-25) was seen to be greatly improved. In one study (Wang et al., 2012), remediating pollutants from the soil by performing tests with salty soil infected with PAH and aryl hydrocarbons was tried. For this analysis, a U-shaped MFC was used, in which simultaneous power generation and removal of hydrocarbons was achieved. Since then, several various forms of BES have shown the probability of effective biodegradation of hydrocarbons, like recalcitrant PAHs, by using many forms of inoculate and hydrocarbon contaminants. In addition it has been shown (Lin et al., 2014) that mixtures of benzene, toluene, ethyl benzene, and xylene (BTEX) degrade in MFC experiments through anoxic oxidation pathways. The MFC network, as the single electron donor, provided a concomitant power production at around 3 mW/m2. It was shown the value of inoculum form in phenanthrene degradation based on MFC by a study (Adelaja et al., 2014) in which Pseudomonas aeruginosa and S. oneidensis were used. High degradation up to 97% was achieved using these strains (Viggi et al., 2017).

4.4.2 Treatment of inorganic pollutants 4.4.2.1 Remediation of bromate and chlorate These inorganic pollutants (chlorate, brominate, etc.) are an increasing concern for the environment and they are present in large quantities in freshwater sources

4.5 Sustainability of the technology

(Hunter, 2002; Butler et al., 2010). These compounds are highly known and also act as carcinogens (Zhao et al., 2012). Therefore, the systems that can transform such toxins successfully are of considerable importance. In a previous insightful study (Adelaja et al., 2017), a tabular up-flow system of MFC was used to concurrently treat these pollutants. Internal condensed catholyte used bromate as the acceptor of electrons. It was known that bromate has high redox potential and it is able to quickly absorb electrons and reduce to bromide ions (Br-1) (Adelaja et al., 2017). The machine was able to extract hydrocarbons in the anode by more than 90%, while the parallel elimination of bromate in the catholyte reached 79%. The elimination in the same method of all these forms of toxins, though generating a power production of about 7 mW/m2, is an important consequence of this research. Bioelectrochemically, perchlorate ions were decreased with a denitrifying biocathode BES shown in a study (Butler et al., 2010).

4.4.2.2 Treatment of heavy metals A small work is focused on researching on heavy metals in BES. Heavy metals are known to have significant threats (Thakur et al., 2016) to the ecosystem, which can incorporate in the food chain, and hence it can cause deleterious effects to the plants and animals. It is, therefore, crucial to search for novel methods of immobilizing and detoxifying free heavy metals introduced to the atmosphere because of anthropogenic behavior. Different approaches are used for the treatment of heavy metals, which include different sorbents (Fu and Wang, 2011) and also reduction electrochemically (Fu and Wang, 2011). Reduction of heavy metals by electrochemical means involves large quantities of energy to be used, which increases the cost of this technique. Using sorbents increases the need for frequent use to recycle the sorbent. This may be complicated, and may entail the involvement of numerous chemicals present, which can affect the process of treatment. A review was carried out (Heijne et al., 2010), which demonstrated that on a twochambered MFC device, copper ions may be decreased and retrieved in their metallic shape. These results suggest BES’s positive capacity for immobilization and retrieval of heavy metal ions from polluted ecosystems.

4.5 Sustainability of the technology It is noted that waste can be utilized for energy production and many valuable products can be obtained. Waste treatment and management techniques and methods depend on many factors like cost of technology and environmentally friendly nature (Massoud et al., 2009; Suriyachan et al., 2012). Economic stability, site requirements, and receiving climate are the most significant of all those considerations. The society will be willing to manage the cost of money, running costs, and maintenance. Because biodiversity is encouraged throughout the water cycle, the technology’s duties will be to recover natural soil, and it must not affect the

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environment. Technology must also help in the production of electricity from the waste materials (Meneses et al., 2010). Sustainability of this technology can be achieved by implanting or integration of different technologies as important. The process must be reliable and feasible to use, and it must not affect the local ecosystem. Also, it is important that people should accept the technology easily, and they should see the benefits. Economic viability is the major step, and it must be employed before utilizing any technique. Many different systems are used but bioelectrochemical systems still need further research and it must be further improved via engineering and biological necessary means. Production cost must be minimized, and current output must be increased. Low-cost materials can be used to decrease the cost of technology. However, due to the complex nature of bioelectrochemical systems, it is expected that the cost of this technology will still be higher than conventional treatment technologies. These systems have some advantages over other technologies compared to anaerobic digestion, such as they can be effective on a limited scale.

4.6 Scaling up of the technology In the past, work focus was laid on the designing of BESs for the large-scale usage. Until now, the bulk of the theoretical work has concentrated on laboratory-scale experiments; although very few experiments have discussed the scaling of bioelectrochemical systems from lab scale to the field scale. It is exhibited from the majority of the scaling-up studies that production of power does not increase due to change in the volume of anodic chamber. However, it was observed that when length of the anodic chamber was increased, BES was operated, and the main demerit raised, from thermodynamics, etc., which then decreases power production or allows for further power input for the electrochemical synthesis using an external power source. High overpotential and lower Coulombic performance are variables restricting the capacity of a single BES. The first functional MFC demonstration is attributed to a study (Dekker et al., 2009), in which they produced stacked MFC consisting of four individual anodic chamber volume units with 5 L. This analysis demonstrated power production of about 144 W/m3. Despite this first effort to build a broader MFC device, there are still several obstacles in the scaling up to produce high current density electrode content, and it is important to use less cost materials for the commercial viability of the system.

4.7 Conclusion It is exhibited from this study that BESs hold great potential for bioremediation of wastewaters but also to produce biorenewable energy using microbes as biocatalysts. In many studies it was observed that these systems can treat different

References

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Acknowledgments The authors would like to express gratitude to Hazara University Mansehra for providing platform and for the support given to conduct the research work.

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Energy generation from fish-processing waste using microbial fuel cells

5

A.R. Abdul Syukor1, Suryati Sulaiman1, Jadhav Pramod Chandrakant1, Puranjan Mishra1, Mohd Nasrullah1, Lakhveer Singh2 and A.W. Zularism1 1

Faculty of Civil Engineering Technology, Universiti Malaysia Pahang (UMP), Kuantan, Malaysia 2 Department of Environmental Science, SRM University-AP, Amaravati, India

5.1 Introduction The use of fossil fuels in recent years has accelerated, especially the use of oil and gas, and this has triggered a global energy crisis. Concerns about climate change and increasing global demand for the finite oil and natural gas reserves are intensifying the search for alternatives to fossil fuels. Renewable bioenergy is viewed as one of the alternative ways to overcome those problems (Wang et al., 2020). A technology using microbial fuel cells (MFCs) that convert the energy stored in chemical bonds in organic compounds to electrical energy achieved through the catalytic reactions by microorganisms has generated considerable interests among academic researchers to find the alternative way for electricity generation. According to Singh et al. (2020a,b), MFC is a device that directly converts chemical energy to electricity through catalytic activities of microorganism. Electricity has been generated in MFCs from various organic compounds including carbohydrates, proteins, and fatty acids. It is made up of two compartments, anode and cathode, separated with proton/cation exchange membrane such as Nafions or salt bridge. The membrane or salt bridge allows hydrogen ion generated in the anode compartment to be transferred into the cathode compartment (Miller et al., 2019). Microorganisms oxidize the substrate and produce electrons and protons in the anode chamber of MFC. Electrons collected on the anode are transported to the cathode by an external circuit (Kakarla and Min, 2019; Lee et al., 2019). Meanwhile, protons are transferred through the membrane internally. Electrons and protons are consumed in the cathode compartment by utilizing oxygen from water (Bond, 2003). Besides its potential to generate electricity, MFC also can be used to measure the strength of wastewater. Application of MFC for wastewater treatment could be an alternative to reduce the cost of treatment. These MFCs was performed well for chemical oxygen demand (COD) and biochemical oxygen demand (BOD) removal that demonstrated the effectiveness of this device for Delivering Low-Carbon Biofuels with Bioproduct Recovery. DOI: https://doi.org/10.1016/B978-0-12-821841-9.00009-8 © 2021 Elsevier Inc. All rights reserved.

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wastewater treatment with removal efficiency of about 90% (Kim, 2002). In MFCs, substrate is regarded as one of the most important biological factors affecting electricity generation. A great variety of substrate can be used in MFCs for electricity production ranging from pure compounds to complex mixture of organic matter present in wastewater. So far, the only objective of the various treatment processes is to remove pollutants from waste streams before their safe discharge into the environment (Liu, 2004). The energy current that is produced by MFC are measured in the current density, which is either represented as the current generated per unit area of the anode surface area (mA/cm2) or current generated per unit volume of the cell (mA/m3). In this study, fresh market waste, which comes from fish, is used as the substrate to generate electricity using MFC technology. Energy has always played an important role in human and economic development and in society’s well-being. For example, fuel wood has been used from time immemorial to make fire, and the first civilizations were already making use of wind in sailing overseas. Without the heat and electricity from fuel combustion, economic activity would be limited and restrained. Modern society uses more and more energy for industry, services, homes, and transport (Glenk and Reichelstein, 2019). This is particularly true for oil, which has become the most traded commodity, and part of economic growth is linked to its price. However, neither oil nor any of the other fossil fuels, such as coal and natural gas, are unlimited resources. The combined effect of growing demand and depleting resources calls for a close monitoring of the energy situation. A shortage of energy sources is not the only problem that happens in this country. Contaminated water or water pollution are also one of the biggest problems that we face. Wastewater from fresh market activity instances, such as vegetable, fish, chicken, and meat processing, has always been neglected. The waste is trapped inside the sump, which causes a strong, disgusting smell, and very disturbing view. Improper management systems will cause the wastewater to go directly to the river or into the runoff system of the area and become the point source of water pollution. Increased awareness among the public and corporate bodies have envisioned the idea of managing the environment in a friendlier manner. It is essential that alternative options are explored to divert and reduce the 670 tons of organic wet market waste from municipal solid waste (MSW). In line with the concept of sustainable development proposed by the United Nations Environment Program (UNEP), more focus had been directed at the possibility of converting waste into value-added products such as generating renewable energy from wet market waste using MFCs. According to estimates done by the International Energy Agency, a 53% increase in global energy consumption is foreseen by 2030, with 70% of the growth in demand coming from developing countries. These growth trends exacerbate the challenge connected to limitations in energy supply, and the resulting competition for resources. Currently, with a total of 20,493 MW installed capacity, the energy reserve margin of Peninsular Malaysia stands at 47%. With an average of 4% annual growth, it is estimated that the maximum demand of electricity will be at 23,099 MW in 2020, which is nearly twice the current demand. Faced with the possibility of a prolonged energy crisis, the government called for the diversification of energy resources other than

5.2 National Green Technology Policy

Table 5.1 Energy mix in Malaysia. Source

1980 (%)

1990 (%)

2000 (%)

2003 (%)

Oil Natural gas Hydro Coal Biomass

87.9 7.5 4.1 0.5

71.4 15.7 5.3 7.6

53.1 37.1 4.4 5.4

6.0 71.0 10 11.9 1.1

crude oil. The Malaysian government also has formulated numerous energy-related policies to ensure the long-term reliability and security of the energy supply for sustainable social-economic development in the country. The National Energy Policy has three primary objectives: supply, utilization, and environmental (Mohamed, 2006). On the other hand, the National Depletion Policy is aimed to conserve the country’s energy resources, particularly oil and gas, as these resources are finite and nonrenewable. In this respect, the production of crude oil was limited to an average of 630,000 barrels per day (bpd) while the consumption of gas in Peninsular Malaysia is limited to about 32,000 million standard cubic feet per day (Mariyappan, 2000). The Fuel Diversification Policy in Malaysia was continuously reviewed to ensure that the country is not too dependent on a single source of energy. Table 5.1 shows the energy mix in Malaysia for the years 1980, 1990, 2000, and 2003.

5.2 National Green Technology Policy To further show the government’s committee to promote low-carbon technology and ensure sustainable development while conserving natural environment and resources, the Malaysian government launched the National Green Technology Policy (NGTP) on July 24, 2009. Green technology refers to the development and application of products, equipment, and systems used to conserve natural environment and resources, which minimizes and reduces the negative impact of human activities and satisfy the following criteria: (1) it minimizes degradation of the environment, (2) it has a zero or low greenhouse gas (GHG) emission, (3) it is safe for use and promotes healthy and improved environment for all forms of life, (4) it conserves the use of energy and natural resources, and (5) it promotes the use of renewable resources. Based on energy trends, the national economic developments are also aligned in the Malaysia Plan (MP). Table 5.2 shows the key emphasis in energy developments from 7 to 10 MP.

5.2.1 Waste from fresh markets The main wet market in Selangor has a daily generation of 15 tons of organic waste, which was disposed of into a sanitary landfill approximately 8 km away from the

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Table 5.2 Malaysia’s key emphasis from 7 to 10 Malaysia Plan for energy development. Malaysia plan

Key emphasis

Seventh Malaysia Plan (1996 2000)

• Emphasis on the sustainable development of depletable resources and the diversification of energy sources.

• Ensuring adequacy of generating capacity as well as

• Eighth Malaysia Plan (2001 05)

•

• • • Ninth Malaysia Plan (2006 10)

•

• •

Tenth Malaysia Plan (2011 15)

• •

expanding and upgrading the transmission and distribution infrastructure. Encouraged the use of new and alternative energy sources as well as efficient utilization of energy. Emphasis on the sustainable development of energy resources, both depletable and renewable. The energy mix includes five fuels: oil, gas, coal, hydro, and renewable energy (RE). Intensify efforts on ensuring adequacy, quality, and security of energy supply. Greater emphasis on energy efficiency (EE). Supports the development of industries in production of energy-related products and services. Incentives for the use of RE resources. Incentives to maintain quality of power supply. Emphasis on strengthening initiatives for EE especially in transport, commercial, and industrial sectors, and in government buildings. Encourage better utilization of RE through diverse fuel sources. Intensify efforts to further reduce the dependency on petroleum, which provides for more efforts to integrate alternative fuels. Incentives in promoting RE and EE are further enhanced. Short term goals vested in NGTP: • Increased public awareness and commitment for the adoption and application of green technology through advocacy programs. • Widespread availability and recognition of green technology in terms of products, appliances, equipment, and systems in the local market through standards, rating, and labeling programs. • Increased foreign and domestic direct investments in green technology manufacturing and services sector. • Expansion of local research institutes and institutions of higher learning to expand research, development, and innovation activities on green technology toward commercialization through appropriate mechanisms. • New RE act mechanism to be launched.

market. When the landfill ceased its operation, the transportation and landfilling costs escalated to approximately USD20 (RM80) per ton from USD10 (RM35). Wet market waste contributes approximately 3.5% of the total waste generated in Malaysia. Increased awareness among the public and corporate bodies has

5.2 National Green Technology Policy

envisioned the idea of managing the environment in a friendlier manner. It is essential that alternative options are explored to divert and reduce the 670 tons of organic market waste from MSW. In line with the concept of sustainable development proposed by UNEP, more focus had been directed at the possibility of converting waste into value-added products, such as organic fertilizer through composting, which requires simple and cheap technologies. Composting is a viable option where it provides various positive outcomes including the simplicity in technology, environmental friendliness, cost, and many other benefits (Tchobanoglous, 1993). Compost enhances the growth of plants with the slow release of necessary minerals and improves the soil condition. Composting has been increasingly popular as an alternative to dispose waste in recent years and a benefiting waste recycling option. Composting reduces and stabilizes the waste and converts it into hygienic and safe products, which add economic value to the final product.

5.2.2 Fish-processing wastewater characteristics Fish-processing wastewater characteristics that raise concern include pollutant parameters, sources of process waste, and types of waste. In general, fishprocessing wastewater can be characterized by its physicochemical parameters, organics, nitrogen, and phosphorus content. Important pollutant parameters of the wastewater are five-day biochemical oxygen demand (BOD5), COD, total suspended solids (TSS), fats, oils, and greases. As in most industrial wastewaters, the contaminants present in fish-processing wastewater are an undefined mixture of substances, mostly organic in nature. Table 5.3 shows the different fish types and their physiochemical parameters.

5.2.2.1 Physiochemical parameters 5.2.2.1.1 pH pH serves as one of the important parameters because it may reveal contamination of a wastewater or indicate the need for pH adjustment for biological treatment of the wastewater. Effluent pH from fish processing is usually close to neutral. The Table 5.3 Raw wastewater characteristic of fish processing industries.

Effluent Flow (L/d) Farmraised catfish Tuna Salmon

79.5K 170K

Five-day biochemical oxygen demand (mg/L)

Chemical oxygen demand (mg/L)

Total suspended solids (mg/L)

Oil and grease (mg/L)

340

700

400

200

1600 300 5500

500 120 1400

250 20 550

246K 13.6M 700 220K 1892.5K 253 2600

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pH levels generally reflect the composition of proteinaceous matter and emission of ammonia compounds.

5.2.2.1.2 Solids content Solids content in wastewater can be divided into dissolved and suspended solids. However, suspended solids are the primary concern since they are objectionable on several grounds. Settleable solids may cause reduction of the wastewater duct capacity (Wang, 2006). When the solids settle in the receiving water body, they may affect the bottom-dwelling flora and the food chain. When they float, they may affect the aquatic life by reducing the amount of light that enters the water. Soluble solids are generally not inspected even though they are significant in effluent with a low degree of contamination. They depend not only on the degree of contamination but also on the quality of the supply water used for the treatment (Wang, 2006). In one analysis of fish filleting wastewater, it is found that 65% of the total solids present in the effluent were already in the supply water (Gonzalez, 2002).

5.2.2.1.3 Odor In fish-processing industries, odor is caused by the decomposition of the organic matter, which emits volatile amines, diamines, and sometimes ammonia. In wastewater that has become septic, the characteristic odor of hydrogen sulfide may also develop. Odor is a very important issue in relation to public perception and acceptance of any wastewater treatment plant. Although relatively harmless, it may affect general public life by inducing stress and sickness (Raychaudhuri and Behera, 2020).

5.2.2.1.4 Temperature To avoid affecting the quality of aquatic life, the temperature of the receiving water body must be controlled. The ambient temperature of the receiving water body must not be increased by more than 2 C 3 C or else it may reduce the dissolved oxygen level. Except for wastewaters from cooking and sterilizing process in canning factories, fisheries do not discharge wastewaters above ambient temperature. Therefore, wastewaters from canning operations should be cooled if the receiving water body is not large enough to restrict the change in temperature to 3 C.

5.2.2.1.5 Organic content The major types of wastes found in fish-processing wastewaters are blood, offal products, viscera, fins, fish heads, shells, skins, and meat (Royintarat et al., 2020). These wastes contribute significantly to the suspended solids concentration of the waste stream. However, most of the solids can be removed from the wastewater and collected for animal food applications. Wastewaters from the production of fish meal, and from solubles and oil from herring, menhaden, and alewives can be divided into two categories: high-volume, low-strength wastes and lowvolume, high-strength wastes.

5.3 Microbial fuel cell system

5.2.2.1.6 Biochemical oxygen demand BOD estimates the degree of contamination by measuring the oxygen required for oxidation of organic matter by aerobic metabolism of the microbial flora. In fishprocessing wastewaters, this oxygen demand originates mainly from two sources. One is the carbonaceous compounds that are used as substrates by the aerobic microorganisms; the other source is the nitrogen-containing compounds that are normally present in fish-processing wastewaters, such as proteins, peptides, and volatile amines. Standard BOD tests are conducted at 5-day incubation for determination of BOD5 concentrations. Wastewaters from fish-processing operations can be very high in BOD5. Literature data for fish-processing operations show a BOD5 production of one to 72.5 kg of BOD5 per ton of product. The BOD is generated primarily from the butchering process and from general cleaning, while nitrogen originates predominantly from blood in wastewater stream.

5.2.2.1.7 Chemical oxygen demand Another alternative for measuring the organic content of wastewater is COD, an important pollutant parameter for the seafood industry. Depending on the type of fish processing, the COD of the wastewater can range from 150 to about 42,000 mg/L. One study examined a tuna-canning and byproduct plant for five days and observed that the average daily COD ranged from 1300 to 3250 mg/L.

5.2.2.1.8 Nitrogen and phosphorus Nitrogen and phosphorus are nutrients that are of environmental concern. They may cause proliferation of algae and affect the aquatic life in a water body if they are present in excess. However, their concentration in the fish-processing wastewater in minimal in most cases. It recommended that a ratio of N to P of 5:1 be achieved for proper growth of the biomass in the biological treatment (Metcalf and Eddy, 2000). Sometimes the concentration of nitrogen may also be high in fish-processing wastewater. One study shows that high nitrogen levels are likely due to the high protein content (15% 20% of wet weight) of fish (Sikrorski, 1999).

5.3 Microbial fuel cell system A typical MFC consists of two separate chambers that can be inoculated with any type of liquid media. These chambers, an anaerobic anode chamber and an aerobic cathode chamber, are generally separated by a proton exchange membrane (PEM) such as Nafion (Kumar et al., 2018). An MFC such as this can be classified into two types. One type generates electricity from the addition of artificial electron shuttles (mediators) to accomplish electron transfer to the electrode. According to Bond (2003), the other type does not require these additions of exogenous chemicals and can be loosely defined as a mediatorless MFC. Mediatorless MFCs can be considered to have more commercial potential than

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MFCs that require mediators because the typical mediators are expensive and toxic to the microorganisms. However, one major disadvantage of the twochamber system in Fig. 5.1 is that the cathode chamber needs to be filled with a solution and aerated to provide oxygen to the cathode (Ren et al., 2019). This same principle can be used to design a single-chamber MFC where the anode chamber is separated from the air-cathode chamber by a gas diffusion layer for a passive oxygen transfer to the cathode, eliminating the need for energy intensive air sparging of the liquid. Meanwhile, according to Du (2007), in the singlechamber MFC, the cathode is exposed directly to the air. The basic component of the MFC is described in Table 5.4.

FIGURE 5.1 Two-chamber MFC with modes of electron transfer (Pant et al., 2016). MFC, Microbial fuel cell.

Table 5.4 Basic components of MFC. Items

Materials

Anode electrode

Graphite, graphite felt, carbon paper, carbon-cloth, Pt, Pt black, reticulated vitreous carbon (RVC) Graphite, graphite felt, carbon paper, carbon-cloth, Pt, Pt black, RVC Glass, polycarbonate, Plexiglas

Cathode electrode Anodic and cathodic chamber proton exchange membrane system (PEM) Electrode catalyst

PEM: Nafion, Ultrex, polyethylene, poly(styrene-codivinylbenzene), salt bridge, porcelain septum, or solely electrolyte Pt, Pt black, MnO2, Fe31, polyaniline, electron mediator immobilized on anode

5.3 Microbial fuel cell system

5.3.1 Substrates used in microbial fuel cell In MFCs, the substrate is regarded as one of the most important biological factors affecting electricity generation (Kumar et al., 2015). A great variety of substrates can be used in MFCs for electricity production, ranging from pure compounds to complex mixtures of organic matter present in wastewater. So far, the only objective of the various treatment processes is to remove pollutants from waste streams before their safe discharge to the environment. In the last century, activated sludge process has been the mainstay of wastewater treatment. However, it is a very energy intensive process and according to an estimate, the amount of electricity needed to provide oxygen in activated sludge process in the United States is equivalent to almost 2% of the total US electricity consumption.

5.3.2 Fish-processing waste as substrate Fish is a very important source of animal protein in the diets of humans. The nutritional properties of fish and fish products render them valuable foodstuffs that are beneficial for human health. Oil-rich fish such as salmon, mackerel, herring, tuna, trout, and sardines are all rich in Omega-3 fatty acids, as shown in Table 5.5 (Usydus et al., 2011). According to Ghassem (2009), Malaysian freshwater fish such as Clarias batrachus (keli) had higher moisture content compared to other freshwater fish. The total lipid content for C. batrachus was approximately ranging from 2.3% to 3.5%, while Pangasius sutchi (patin) fat content was 8.8%. The crude protein content ranged from 13.3% in P. sutchi to 21% in Channa striatus (haruan) and Oreochromis niloticus (tilapia). The major amino acids in all four fish were

Table 5.5 Fish species and their Omega-3 fatty acid content. Fish species

Omega-3 content (g) per 100 g of fish

Tuna (fresh) Atlantic salmon Mackerel Atlantic herring Rainbow trout Sardines Halibut Tuna (canned) Cod Haddock Catfish

0.28 1.51 1.28 2.15 0.4 1.85 2.01 1.15 1.25 2 0.47 1.18 0.31 0.28 0.24 0.18

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glutamic acid, aspartic acid and lysine, ranging from 3.8% to 13.1% of total amino acids. The results showed that freshwater fish could be used as a major source for healthy diets based on protein content as well as high amino acid and fatty acid composition. It is also suitable for use as a substrate in MFC due to its level of fatty acids (Fregui, 2010).

5.4 Treatment methodology of fish-waste using microbial fuel cell (a Malaysian case study) The study area for this research only focuses on one location, which is Pasar Besar Kuantan, Kuantan, Pahang. There are a few reasons why this place was selected as the study area. The first reason is this market is the main and the largest wet market in Kuantan. It is located in the city center of Kuantan. Most of the fish are taken fresh from the sea before been iced, and they make their way to the other urban centers in the peninsula. Second, because this market was built more that 20 years ago, there is no proper drainage for the wastewater. The wastewater, whether from fish, fruit, or vegetable stalls, are going directly to the nearest drainage around the market. Then, it will discharge into the nearest river every day. This creates a pollution problem in the Kuantan river. Meanwhile, the solid waste from the market are directly dumped and disposed of at the Kuantan Jabor-Jerangau landfill, and the cost for disposal and maintenance of the solid waste from the market only takes more than RM 1 million a year.

5.4.1 Preparing the substrate In this study, the waste from processing catfish and shark catfish, which is the tail, head, skin, intestinal organ, and blood, are collected from the marketplace. This is shown in Figs. 5.2 and 5.3. The waste is then divided into three compositions, which is three-kilogram waste from catfish processing, threekilogram waste from shark catfish processing, and another three-kilogram waste from a combination of both fish. All three-composition waste is blended using a blender, which is shown in Fig. 5.4.

5.4.2 Testing for physical, chemical, and biological parameters After the waste is blended, it will be tested for physical, chemical, and biological characteristics. Most of the testing is carried out by using Spectrophotometer DR/4000 that is based on SMEWW, which is shown in Table 5.6. There are eight parameters tested in this study. Those parameters are pH, BOD, COD, oil and grease, temperature, nitrogen, and phosphorus.

5.4 Treatment methodology of fish-waste using microbial fuel cell

FIGURE 5.2 Flowchart of experimental design.

5.4.3 Electrode The electrodes that have been used in this MFC are made from graphite. The graphite is cut into two sizes, 60 3 60 and 110 3 50, which are shown in Fig. 5.5. Then, two holes are drilled on the graphite plate. This hole is used to insert two

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FIGURE 5.3 Heads, skins, tails, and intestinal organs of fish.

FIGURE 5.4 The fish waste was blended and weighed according to the composition.

Table 5.6 List of physical, chemical, and biological parameters. No

Parameter

Standard code APHA

1 2 3 4 5 6 7 8

Biochemical oxygen demand (BOD) Chemical oxygen demand (COD) Total organic carbon (TOC) Temperature Oil and grease Nitrogen Phosphorous pH

APHA 5210 B# APHA 5220 B# APHA 5310 B# APHA 5210 B# APHA 5520 B# APHA 4500-NO2 B# APHA 4500-PB4 B# APHA 4500-H1-B (c)#

# APHA

5.5 Results observation

FIGURE 5.5 The schematic diagram of a small electrode.

stainless steel screw holders so that the electrode is in a hanging position. Then, the electrode is tied to the container lid using a bolt and nut.

5.5 Results observation The results gathered from the experiments are analyzed by using Statistical Package for the Social Science (SPSS) Statistic 17.0 software through One-way Analysis of Variance (ANOVA). One-way ANOVA is a technique used to compare means of two or more samples (using the F distribution). The ANOVA tests the null hypothesis that samples in two or more groups are drawn from the same population. To do this, two estimates are made of the population variance. These estimates rely on various assumptions. The ANOVA produces an F statistic, the ratio of the variance calculated among the means to the variance within the samples. If the group means are drawn from the same population, the variance between the group means should be lower than the variance of the samples. A higher ratio therefore implies that the samples were drawn from different populations.

5.5.1 Voltage production All voltage results are obtained with MFC reactors at laboratory room temperature, normally 20 C during daytime. The voltage readings are recorded for 30 days at the same time each day. The summary of voltage according to type of substrate and size of electrode are shown in Fig. 5.6.

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FIGURE 5.6 Summary of voltage production based on type of substrate and size of electrode.

Between the three compositions of waste, the shark catfish, using the larger electrode, produced a higher voltage, 0.52 V, compared to other compositions. The same result was also obtained for the smaller electrode and is shown in Fig. 5.7. The MFC that used a small electrode with the shark catfish as substrate produced higher voltage, which is 0.316 V. Shark catfish (P. sutchi) was able to produce more voltage compared to catfish (C. batrachus) and integrated with both fish because it has high Omega-3 and Omega-6 fatty acid (Usydus et al., 2011). According to Fregui (2010), MFC that use fatty acid as a substrate are able to produce 0.23 V in 30 days of operation. For MFC with the smaller electrode, the same substrate was also managed to produce the maximum voltage. Nevertheless, the voltage production are less than MFCs with larger electrode. This is because the surface area of electrodes will affect the voltage production in the MFC. Wang et al. (2018) have stated that as the area of anode was increased, power density was increased from 0.086 to 0.115 V, demonstrating that power density was not doubled when the second electrode was added.

5.5 Results observation

FIGURE 5.7 Percentage BOD removal for two different size electrodes. BOD, Biochemical oxygen demand.

5.5.2 Biochemical oxygen demand removal BOD is the amount of oxygen consumed by the organism in the process of stabilizing waste. It means that if sufficient oxygen is available, the aerobic biological decomposition of an organic waste will continue until all of the waste is consumed. Based on the comparison between the large and small electrode sizes, it appears that the large electrode removed more BOD that the smaller size. This is shown in Fig. 5.8. MFC that used shark catfish as substrate with the large electrode removed 63.4% of BOD. This shows that different of substrate will affect the percentage of BOD removal. Liu (2004) said that the percentage of BOD removal is based on its substrate. Meanwhile, MFC that used the same substrate but with a smaller electrode only managed to remove 59.3% of BOD. This result was statically tested using ANOVA and according to it, there is significant difference statistically between the types of substrate and size of electrode used in percentage of BOD removal (P 5 .007).

5.5.3 Chemical oxygen demand removal COD is a measure of the capacity of water to consume oxygen during the decomposition of organic matter and the oxidation of inorganic chemicals such as

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FIGURE 5.8 Percentage COD removal for two different size of electrode. COD, Chemical oxygen demand.

ammonia and nitrate. The COD test normally yields higher oxygen-equivalent values than using the standard BOD5 test, because more oxygen equivalents can always be oxidized by the chemical than can be oxidized by microorganisms. The best COD removal were obtained in MFC that used shark catfish as substrate with the larger electrode. Those MFC have removed 68.9% of COD in a month. Meanwhile, MFC that used the smaller electrode only removed 65.5% of COD with shark catfish as the substrate. With ANOVA analysis, it proves that there is significant difference statistically between the types of substrate and size of electrode used in percentage of COD removal (P 5 .01).

5.5.4 Nitrogen Before inserting the waste into MFC, there is no nitrogen content in all substrate as shown in Fig. 5.9. Nevertheless, while the MFC is running, there is an increase of nitrogen content in all substrates. This is because, nitrogen gas has been inserted inside the MFC to remove the oxygen content. The highest nitrogen content is MFC with catfish as a substrate using the smaller electrode. The value is 16,320 mg/L, and it was obtained while working the MFC.

5.5 Results observation

FIGURE 5.9 Total nitrogen content between two sizes of electrode.

Table 5.7 One-way Analysis of Variance table for nitrogen by type of substrates.

Between groups Within groups Total

Sum of squares

df

Mean square

F

Sig.

3.047E7 7.305E8 7.609E8

2 33 35

1.524E7 2.214E7

0.688

0.510

Meanwhile, after the MFC stopped, the value of nitrogen decreases almost 26% for catfish substrate with the smaller electrode. It also shows that the nitrogen content is decreasing for all types of substrate weather using a larger or smaller electrode. This could be attributed to the ammonia stripping in the cathode compartment due to aeration or simultaneous nitrification and denitrification occurring in the cathode compartment (Ghosh et al., 2020). One-way ANOVA analysis in Table 5.7 shows that P 5 .510 . .05 for nitrogen by type of substrate. Therefore, alternative hypotheses were rejected, and there is no significant different between the types of substrate used in amount of total nitrogen. For nitrogen by size of electrode ANOVA analysis (Table 5.8) the

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Table 5.8 One-way Analysis of Variance table for nitrogen by size of electrode. Between groups Within groups Total

Sum of squares

df

Mean square

F

Sig.

2.969E8 4.640E8 7.609E8

1 34 35

2.969E8 1.365E7

21.757

0.000

FIGURE 5.10 Phosphorus content in MFC. MFC, Microbial fuel cell.

P value is equal to .000, and the null hypothesis was used, which is that there is significant differences statistically between the size of electrode used in the amount of total nitrogen.

5.5.5 Phosphorous Phosphorous is a multivalent nonmetal of the nitrogen group. It is found in nature in several allotropic forms, and is an essential element for the life of organisms. As shown in Fig. 5.10, MFC that used catfish as the substrate have the highest content of phosphorus, which is 39.5 mg/L. This is because catfish has higher

5.6 Conclusion

Table 5.9 One-way Analysis of Variance table for phosphorus by type of substrates. Between groups Within groups Total

Sum of squares

df

Mean square

1019.813 663.188 1683.000

2 33 35

509.906 20.097

F

Sig.

25.373

0.000

phosphorus content compared to other fish. The same result was obtained from MFC that used the smaller electrode. Catfish as a substrate gave more phosphorus content at 38.5 mg/L. One-way ANOVA analysis in Table 5.9 shows that P 5 .000 , .05 for phosphorus by type of substrates. Therefore, the null hypothesis was rejected, and there is significant difference statistically between the types of substrates in the phosphorus content. Meanwhile, in Table 4.10, the P value is bigger than .05. Therefore, there is no significant difference between the types of substrates in the phosphorus content.

5.6 Conclusion Overall, fish-processing waste, which comes from C. batrachus (catfish) and P. sutchi (shark catfish), is capable for use as a substrate in an MFC. Those wastes are able to produce electricity and simultaneously able to remove more than 60% of BOD and COD. It also shows that more voltage will be generated if the surface area of the electrodes are increased. In MFCs that used larger electrodes, P. sutchi processing waste is able to produce 0.508 V within a month compared to C. batrachus, which only produced 0.42 V. Meanwhile, the combination of both wastes were only capable to produce a maximum of 0.27 V. The same thing happens in MFCs that use smaller electrodes. P. sutchi waste was able to produce the highest maximum voltage compared to other wastes, but it cannot produce voltage more than 0.37 V. This happens because of its limitations from a smaller surface area of electrodes. By using ANOVA with a significant level of 0.05, it has been proved statistically that there is a significant difference between the types of substrate used in producing more voltage (P , .05). It also proved that there is significant difference statistically between sizes of electrodes used in producing more voltage. The ability to remove BOD and COD is one of the advantages of using MFC as a technology in generating alternative energy. MFCs are not only capable of generating energy, but they also can remove BOD and COD simultaneously. This advantage is very useful because most of the substrates that have been used in MFC have contained high BOD and COD concentrations. Wastes that have high BOD and COD have more treatment costs if they go directly inside a treatment plant. By using MFC, some of the BOD and COD are able to be removed, therefore reducing the treatment costs. In this study, by using a large electrode, the MFC was able to

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CHAPTER 5 Energy generation from fish-processing waste

remove 63.4% of BOD and 68.9% of COD. Meanwhile, MFCs that use smaller electrodes were able to remove 59.3% of BOD and 65.5% of COD inside the waste. With this result, it can be concluded that increasing the size of the electrode used in MFC also can increase the percentage of BOD and COD removal.

References Bond, D.L., 2003. Cathode performance as a factor in electricity generation in microbial fuel cells. Appl. Environ. Microbial. 38, 1548 1555. Du, Z.L., 2007. A state of the art revew on microbial fuel cells; a promising technology for wastewater treatment and bioenergy. Biotech. Adv. 25, 464 482. Fregui, S.E., 2010. Microbial fuel cells operating on mixed fatty acids. Bioresources Tech 101, 1233 1238. Ghassem, M.K., 2009. Proxiamte composition, fatty acid and amino acid profiles of selected Malaysian freshwater fish. Malays. Fish. J. 8, 7 16. Ghosh, P., Kumar, M., Kapoor, R., Kumar, S.S., Singh, L., Vijay, V., et al., 2020. Enhanced biogas production from municipal solid waste via co-digestion with sewage sludge and metabolic pathway analysis. Biores. Technol. 296, 122275. Glenk, G., Reichelstein, S., 2019. Economics of converting renewable power to hydrogen. Nature Energy 4 (3), 216 222. Available from: https://doi.org/10.1038/ s41560-019-0326-1. Gonzalez, J.C., 2002. Composition of fish filleting wastewater. Environ. Technol. Lett. 269 272. Kumar, R., Singh, L., Wahid, Z.A., Din, M.F.M., 2015. Exoelectrogens in microbial fuel cells toward bioelectricity generation: a review. Int. J. Energy Res. 39 (8), 1048 1067. Kumar, R., Singh, L., Zularisam, A.W., Hai, F.I., 2018. Microbial fuel cell is emerging as a versatile technology: a review on its possible applications, challenges and strategies to improve the performances. Int. J. Energy Res. 42 (2), 369 394. Kakarla, R., Min, B., 2019. Sustainable electricity generation and ammonium removal by microbial fuel cell with a microalgae assisted cathode at various environmental conditions. Bioresour. Technol. 284, 161 167. Available from: https://doi.org/10.1016/j. biortech.2019.03.111. Kim, H.P., 2002. A mediator-less microbial fuel cell using a metal reducing bacterium, Shewanella putrefaciens. Enzyme Microb. Technol. 145 152. Lee, M., Kakarla, R., Min, B., 2019. Performance of an air-cathode microbial fuel cell under varied relative humidity conditions in the cathode chamber. Bioprocess Biosyst. Eng. 42 (8), 1247 1254. Available from: https://doi.org/10.1007/s00449-019-02122-9. Liu, H.L., 2004. Electricity generation using an air-cathode single chamber microbial fuel cell in the presence and absence of a proton exchange membrane. Environ. Sci. Technol 38 (14), 4040 4046. Available from: https://doi.org/10.1021/es0499344. Mariyappan, K., 2000. Retrieved March 14, 2010, from Institute for Sustainable Energy Policies: www.isep.or.jp/spena/2000/countryreports/malaysia.htm. 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. Metcalf & Eddy, 2000. Wastewater Engineering: Treatment, Disposal, Reuse. McGrawHill Book Co, New York.

References

Mohamed, A.K., 2006. Energy for sustainable development in Malaysia: energy policy and alternative energy. Energy Policy 34 (15), 2388 2397. Raychaudhuri, A., Behera, M., 2020. Comparative evaluation of methanogenesis suppression methods in microbial fuel cell during rice mill wastewater treatment. Environ. Technol. Innov. 17, 100509. Available from: https://doi.org/10.1016/j.eti.2019.100509. Ren, P., Ci, S., Ding, Y., Wen, Z., 2019. Molten-salt-mediated synthesis of porous Fecontaining N-doped carbon as efficient cathode catalysts for microbial fuel cells. Appl. Surf. Sci. 481, 1206 1212. Available from: https://doi.org/10.1016/j.apsusc.2019.03.279. Royintarat, T., Choi, E.H., Boonyawan, D., Seesuriyachan, P., Wattanutchariya, W., 2020. Chemical-free and synergistic interaction of ultrasound combined with plasmaactivated water (PAW) to enhance microbial inactivation in chicken meat and skin. Sci. Rep. 10 (1), 1 14. Available from: https://doi.org/10.1038/s41598-020-58199-w. Singh, L., Rana, S., Thakur, S., Pant, D., 2020a. Bioelectrofuel synthesis by nanoenzymes: novel alternatives to conventional enzymes. Trends Biotechnol. 38, 469 473. Singh, L., Mahapatra, D.M., Liu, H., (Eds.), 2020b. Novel Catalyst Materials for Bioelectrochemical Systems: Fundamentals and Applications. American Chemical Society. Sikrorski, Z., 1999. Seafood Resource: Nutritional Composition and Preservation. CRC Press. Inc, Boca Raton, FL. Tchobanoglous, G.T., 1993. Integrated Solid Waste Management: Engineering Principle and Mangement Issue. McGraw Hill Inc, New York. Usydus, Z., Szlinder-Richert, J., Adamczyk, M., Szatkowska, U., 2011. Marine and farmed fish in the Polish market: comparison of the nutritional value. Food Chem. 126 (1), 78 84. Wang, K.H., 2006. Waste Treatment in The Food Processing Industry. Taylor & Francis. Wang, L., Singh, L., Liu, H., 2018. Revealing the impact of hydrogen productionconsumption loop against efficient hydrogen recovery in single chamber microbial electrolysis cells (MECs). Int. J. Hydrogen Energy 43 (29), 13064 13071. Wang, L., Chen, Y., Long, F., Singh, L., Trujillo, S., Xiao, X., et al., 2020. Breaking the loop: tackling homoacetogenesis by chloroform to halt hydrogen productionconsumption loop in single chamber microbial electrolysis cells. Chem. Eng. J. 389, 124436.

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CHAPTER

Microbial electrosynthesis: Recovery of high-value volatile fatty acids from CO2

6

Narnepati Krishna Chaitanya, Akash Tripathi and Pritha Chatterjee Department of Civil Engineering, Indian Institute of Technology Hyderabad, Hyderabad, India

6.1 Introduction Fossil fuels have been intensively used for energy production, causing a dramatic increase of CO2 levels in the atmosphere and related environmental impacts, such as global warming and extreme weather events. To counteract those issues, in 2015, 195 countries signed the Paris Agreement and committed to cut off their greenhouse gases emission by 40% by 2030 (Liobikiene
and Butkus, 2017). The achievement of such ambitious targets is linked to a gradual switch from fossil fuels to sustainable and renewable sources of energy for production of energy and chemical commodities. This is driving carbon-intensive industries such as paper, food, energy, cement, and petroleum-refining industries toward a circular economy concept, in which side and waste streams, including gaseous CO2 streams, are seen as a potential feedstock for fuel and/or chemical production. Indeed, CO2 is a building block for synthesizing a wide array of chemical and energy products through either chemical or biotechnological routes. Biological CO2 utilization processes, in which enzymes or microorganisms are employed to produce useful products from CO2, are inherently circular, enable carbon-neutral or even carbonnegative balance, and are less energy demanding than chemical processes (Aresta et al., 2016). Microbial electrosynthesis of high-value compounds from CO2 is an emerging area of research and is one of the most potential technologies coupling reduction of carbon emissions to the biological production of chemical commodities. In microbial electrosynthesis cell (MES), electroactive microbes can use electrons directly or indirectly with help of mediators from a cathode electrode to reduce CO2 into value-added organic compounds (Nevin et al., 2011). Since Nevin et al. (2010) investigated the first proof-of-concept, acetate was produced as the main compound from MES using CO2 (Batlle-Vilanova et al., 2017). Several studies have focused on enhancement of MES performance in-terms of acetate production (Jourdin et al., 2016; LaBelle and May, 2017). Effects of electrode materials and designs (Bajracharya et al., 2015; Song et al., 2018; Jourdin et al., 2014, 2015; Chen et al., 2016; Marshall et al., 2013; Nie et al., 2013; Zhang et al., 2013), impact of inoculum (Blanchet et al. 2015; Mateos et al. 2018; Delivering Low-Carbon Biofuels with Bioproduct Recovery. DOI: https://doi.org/10.1016/B978-0-12-821841-9.00006-2 © 2021 Elsevier Inc. All rights reserved.

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Tremblay et al., 2015), influence of power fluctuations (Mateos et al., 2020; del Pilar Anzola Rojas et al., 2018a,b), impact of operational conditions such as pH (Batlle-Vilanova et al., 2016; Chen et al., 2015; Jourdin et al., 2016; Marshall et al., 2013), applied potentials (Ameen et al., 2020; Blanchet et al. 2015; Irfan et al., 2019; Jourdin et al., 2016; Mateos et al., 2018; Mohanakrishna et al., 2016), and reactor design (Bajracharya et al., 2016; Giddings et al., 2015; Jourdin et al., 2018; Srikanth et al., 2018a) on MES performance have been extensively studied. There are many reviews published on MES in different aspects for example on reactor design (Roy et al., 2015), electron transfer mechanism (May et al., 2016), biocatalysts used (Bajracharya et al., 2017a,b), and electrode materials (Aryal et al., 2017a). However, there is no comprehensive review on the various factor affecting the performance of MES and significance of each individual factor. A critical analysis of the parameters affecting product synthesis in MES is presented in this chapter with a statistical analysis of the performance. An economic analysis of microbial electrochemical systems is also done. This book chapter could provide an insight into optimization studies for MES reactor and suggest helpful insights into the commercial application of MES in large-scale systems to enhance the efficiency and economic value of chemical synthesis.

6.2 Basic principle of microbial electrosynthesis cell Bioelectrochemical system (BES) is a microbially assisted electrochemical technology, which has two half cells that are alienated by a proton exchange membrane (Liu and Logan, 2004; Rozendal et al., 2006). These processes produce electricity from organic matter or high-value chemicals from waste/wastewater. At the anode electrode in a BES, chemicals and organic matter are oxidized by microorganisms to generate protons, and the electrons are released. These electrons travel to the cathode electrode by a conductive circuit. Protons formed in the reaction are proficient to travel (diffusion of protons is due to concentration gradient) in an electrolyte (Gil et al., 2003). These protons combine with electrons at the cathode electrode to produce H2 gas (Fig. 6.1). The produced hydrogen then reduces CO2 at the cathode of a MES. This last part of synthesis of chemicals need an additional power supply to drive the reaction due to its nonspontaneous reaction nature.

6.3 Factors affecting product titer Electrotrophs consume electrons from electrodes and can provide a broad range of functions in MES. Their outputs depend on the microorganisms present, microbial activity, and the electrode potential. While identification of these microorganisms and their mechanisms of electron uptake are still developing, there has been

6.3 Factors affecting product titer

FIGURE 6.1 Mechanism of electron transfer and chemical synthesis in BES. BES, Bioelectrochemical system; VFA, volatile fatty acids.

considerable research so far. The performance of MES systems depends on a number of factors, such as, pH, electrode potential, fluctuations in power supply, electrode material, and inoculum.

6.3.1 The effect of pH Every living cell has maximum activity in a specific range of pH. Similarly, chemical reactions also are often pH dependent. Since the reactions in a MES are bioelectrochemical, it is evident that pH of anolyte or catholyte has impact on the rate of production of volatile fatty acids (VFAs) (Batlle-Vilanova et al., 2016; Grethlein et al., 1990). In most studies, it was perceived that slightly acidic conditions (5.226.7) are favorable for VFA production and it controls methanogenic bacteria as well (Jourdin et al., 2016; Mateos et al., 2018). A low pH (,4.5) and a high pH ( . 9) can cause harm to acetogenic microorganisms and can also prevent biofilm development on cathode electrodes (Jourdin et al., 2016; Mohanakrishna et al., 2015). Jourdin et al. (2016) observed an acetate production rate of 1330 g/m2/d at pH 5.2 and reported that higher proton availability drastically increases acetate production rate. When CO2 is fed to the cathode chamber it dissolves in the catholyte and forms bicarbonates in the system which increases the pH of the catholyte. LaBelle et al. (2014) observed that acetate production rate dropped from 51.6 to 0 mM/d when the media pH changed from 7 to 5.5.

125

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CHAPTER 6 Microbial electrosynthesis: Recovery

With decreasing catholyte pH, concentration of both CO2 and protons increases, so that acetate production became thermodynamically favorable. In a MES, acetate can be produced by two mechanisms: (1) bioelectrochemically (Eq. 6.1), and (2) through hydrogen-mediated mechanism (Eqs. 6.2 and 6.3).
CO2 17H1 18e2 -CH3 COO2 12H2 O; Eo 5 2 280 mV versus SHE Ag=AgCl 1

2

(6.1)

2H 12e -H2

(6.2)

CO2 14H2 -CH3 COO2 1H1 12H2 O

(6.3)

At lower pH, according to Nernst equation, hydrogen electrolytic reduction could also become thermodynamically favorable due to the higher proton availability, thus increasing the production rate of acetate. Other than that, due to less pH and higher substrate availability, some autotrophic microorganisms become more active. For example, microorganisms like Clostridium ljungdahlii can obtain energy from the proton gradient between inside and outside the cell and drive the metabolism, thus increasing the acetate production rate (Batlle-Vilanova et al., 2016).

6.3.2 Fluctuations in electricity supply Fluctuations in electricity or break in supply at different intervals can cause change in microbial activity and affect the performance of MES. del Pilar Anzola Rojas et al. (2018a) disconnected MES reactor from power supply for 4, 6, 8, 16, 32, and 64 hours and observed acetate production rates. Interruptions affected the acetate production rate, causing a decrease of 77% after 64 hour off. However, after all the interruptions, the acetate production was restored, taking between 7 and 16 hours for the reduction current to turn steady with longer recovery times for longer off periods. The microbial community on MES proved to be resilient and able to recover the electroautotrophic activity despite the duration of current supply interruptions. del Pilar Anzola Rojas et al. (2018b) further observed in a similar study that during the absence of power supply, it was observed that acetic acid is oxidized back to CO2, which suggests microbial activity and/or pathway reversal. After reaching optimal acetate yield in a MES, the cell was left in open circuit for six weeks, not adding any substrate in the medium or gases in the headspace by Mateos et al. (2020). Acetogenic and H2-producing bacteria activity recovered after reconnection. However, few days later syntrophic acetate oxidation bacteria and H2-consuming methanogens became dominant, producing CH4 as the main product, via electromethanogenesis and the syntrophic interaction between eubacterial and archaeal communities which consume both the acetic acid and the hydrogen present in the cathode environment. These studies reveal that though bacterial population have an ability to recover from power fluctuations, but only up to a certain extent and prolonged duration of power cut can cause irreversible damage to microbial community. Resilience of the microbial community to fluctuations in power supply shows the potential of coupling MES

6.3 Factors affecting product titer

technology with renewable energy sources, although extensive further work must be carried out to achieve real-field implementation of MES systems together with renewable energy power plants. When intermittent energy sources are used to power MES, additional investment is required for energy storage systems to recover excess unused energy from the production peaks, and deliver it to the MES reactors with a constant flow.

6.3.3 Impact of inoculum The choice of inoculum depends on the purpose of the biocathode or the expected product. For acetate bioproduction processes starting from CO2, homoacetogens are the obvious choice of microorganisms because of their ability to fix CO2 via the efficient Wood Ljungdahl pathway. Only a few pure strains, for example Sporomusa ovata (Aryal et al., 2016, 2017b; Blanchet et al. 2015; Nevin et al., 2010; Nie et al., 2013; Zhang et al., 2013), C. ljungdahlii (Bajracharya et al., 2015), Sporomusa silvacetica, Sporomusa sphaeroides, Clostridium aceticum, and Moorella thermoacetica (Lovley et al., 2011) have been reported to use electricity for converting CO2 into valuable chemicals such as acetate. While acetate is the typical end product for all acetogens, butyrate is one of the end products of Clostridium carboxidivorans/Clostridium ragsdalei (Daniell et al., 2012; Ganigue´ et al., 2015). Mixed microbial cultures offer advantages over pure cultures such as the possibility to operate reactors without sterilization, and resistance to operational perturbations. However, when using mixed cultures, the cathodic product titer is often low due to competitive metabolic pathways, especially methanogenesis (Chatterjee et al., 2019). To overcome this researchers often inoculate biocathodes with an already acclimated microbial community to reduce the start-up time and increase performance. To enrich homoacetogens in a mixed culture the first step followed by researchers is to suppress the methanogens and then later to promote acetogenic growth by growing in a H2:CO2 headspace (Annie Modestra et al., 2015; Mohanakrishna et al., 2015; Patil et al., 2015). Patil et al. (2015) used a mixture of anodic effluent of an MFC and the effluent of a upflow anaerobic sludge blanket reactor (UASB) as source of inoculum for producing butyrate in a MES. The selection of these two inoculum sources was based on the theoretical possibility of the presence of the needed functional traits within the microbial community, that is, interaction with an electrode (anode biomass) and the formation of products from CO2 as a carbon source (UASB biomass). The mixture was dosed with 100 mM 2-bromoethanesulfonicacid (BESA) to reduce methanogens, enriched under a H2:CO2 headspace of 70:30 for 110 days and subsequent quick culture transfers (1000 3 dilution every 2 days for three consecutive cycles) were performed before use in MES. The enriched acetogenic culture grown on H2:CO2 was dominated by Clostridium (Patil et al., 2015). A similar strategy to enrich homoacetogens by suppressing methanogens using BESA and later growing it in H2:CO2 headspace was used by Annie Modestra et al. (2015). Mohanakrishna et al. (2015) used

127

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CHAPTER 6 Microbial electrosynthesis: Recovery

granular sludge from an anaerobic digester and enriched it in steps, by first removing the methanogens by heating the sludge at 90℃ for 1 hour, then culturing it in glucose media to allow heterotrophic growth followed by changing the media to a mixture of H2 and CO2 to ensure autotrophic homoacetogenic growth. Electrical inversion of bioanodes to H2-evolving biocathodes was first demonstrated by Rozendal et al. (2008), who enriched bioanodes on hydrogen to obtain hydrogen-evolving biocathodes based on the reversibility of hydrogenases. Mateos et al. (2018) observed that a nonspecialized inoculum resulted in a highly specialized cathodic biofilm by polarity reversal, which was accompanied by an improved performance in terms of consumed current and product generation. Interestingly, a much more specialized inoculum promoted a rediversification in the biofilm with a lower general cell performance. A higher specialization of the biofilm led to an improvement in acetate generation, probably due to lowered influence of undesirable secondary metabolic pathways (Mateos et al., 2018). Hartline and Call (2016) observed that high-current generating bioanodes do not necessarily translate into high-current consuming biocathodes after polarity reversal.

6.3.4 Electrode materials Electrode material with high surface area supports high biofilm growth, which in turn delivers/consumes current at a faster rate in electrosynthesis processes (Das et al., 2018). Poor electron transfer is observed in electrode materials with lower surface area, which can be improved by facilitating high surface area to volume ratio (Aryal et al., 2017a). Cathode material should have good biocompatibility, low cost, easy availability, high capability of H2 evolution, and at the same time it should show high holding capacity of H2 (Bajracharya et al., 2015). Commercially available electrode materials, such as unpolished graphite sticks, were used in the beginning stage of MES, reported by Nevin et al. (2010). In recent years, materials like granular graphite, carbon felt, carbon cloth, carbon nanotube (CNT), and 3D reticulated vitreous carbon (RVC) electrodes were successfully used as electrode material in MES giving higher acetate production, compared to flat electrodes (Table 6.1) (Ganigue´ et al., 2015; Jourdin et al., 2014, 2018; Marshall et al., 2013). 3D RVC electrodes produced a high current density of 3.7 mA/cm2 and a maximum acetate production of 1.3 mM/cm2/d, producing 1.7 times and 2.6 times higher current density and acetate, respectively, as compared to a flat carbon electrode (Jourdin et al., 2014). Until 2016, multiwalled CNT-RVC used by the same research group recorded the highest acetate production (1330 g/m2/d) (Jourdin et al., 2016). Surface chemistry modifications to the electrodes resulted in better microbeelectrode interactions and resulted in a high rate of acetate production as well as maximum current consumption (Aryal et al., 2017b). Zhang et al. (2013) used a novel cathode electrode, and the surface of carbon cloth was modified by treating it with chitosan; they reported that the modified electrode material showed 7.6-fold

6.3 Factors affecting product titer

Table 6.1 Comparison between surface modified and untreated cathode materials on acetate production in microbial electrosynthesis cell.

Modified surface electrode Chitosan treated carbon cloth Ni-nanowire graphite NanoWeb-reticulated vitreous carbon (RVC) 3D-graphene carbon felt composite Reduced graphene oxide tetraethylene pentamine carbon cloth (rGO-TEPA-CC) Iron oxide modified 3D carbon felt Graphene paper Graphene nickel foam MnO2 coated carbon felt rGO coated copper foam

Titanium carbon coated on carbon felt

Rate of production improvement (acetate)

Compared with unmodified/ untreated

References

7.6-fold

Carbon cloth

Zhang et al. (2013)

2.3-fold 33.3-fold

Graphite electrode Carbon plate

Nie et al. (2013) Jourdin et al. (2014)

6.8-fold

Carbon felt

Aryal et al. (2016)

11.8-fold

Carbon cloth

Chen et al. (2016)

4.8-fold

Carbon felt

Cui et al. (2017)

8-fold 1.8-fold

Carbon paper Nickel foam

Aryal et al. (2017a) Song et al. (2018)

1.7-fold

Carbon felt

Anwer et al. (2019)

21.3-fold 43.5-fold

Copper foam rGO foam electroplated with copper Carbon felt

Aryal et al. (2019)

1.6-fold

Tahir et al. (2020)

(13.51 6 3.30 g/m2/d) higher acetate production compared to unmodified carbon cloth cathode. Treatment with chitosan resulted in biocompatibility of the electrode causing improved electrostatic interactions between negatively charged bacterial cells and the positively charged chitosan-modified electrode. Surface modifications of 3D electrode materials for further improved current density can be considered for future scaling-up studies or for practical field-scale implementation.

6.3.5 Effect of electrode potential For each microorganism there is a particular redox potential where it has maximum electrochemical activity (Kracke et al., 2015). Marshall et al. (2013) used enriched inoculum (from acetogenic biocathode) and poised the cathode at 20.59 V and achieved a maximum production of 17.25 mM/d acetate. A study

129

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CHAPTER 6 Microbial electrosynthesis: Recovery

by Van Eerten-Jansen et al. (2013) showed that at a higher overpotential of 20.9 V, caproate and butyrate were the main products in a MES inoculated with anaerobic sludge. Blanchet et al. (2015) observed that there is no acetate production at high negative overpotentials (20.36 V), whereas low reduction potentials (20.66 V) enabled H2 evolution and acetate production up to 244 6 20 mg/L. Mohanakrishna et al. (2016) studied the effect of cathode potentials at 20.6 and 20.8 V and observed higher coulombic efficiency at 20.8 V. However, they observed potential losses that are also very high at higher cathodic overpotential when compared to low applied potentials. The reduction potential to drive the cathode reaction should be in the range of 20.4 to 21.4 V. Nie et al. (2013) observed that at an imposed potential of 20.4 V to the cathode, a maximum acetate production rate of 1.13 mM/d at Coulombic efficiency (CE) 86%, could be obtained with S. ovata grown in DSMZ2662. Mohanakrishna et al. (2015) inoculated the cathode, imposed at 20.4 V, with S. ovata grown and achieved a maximum acetate production rate of 2.04 mM/d at 90.61% CE. Overall, a higher overpotential has shown an increased rate of acetate production up to 21.4 V. However, maintaining higher cathode reduction potential can also lead to methane production and increased potential losses (Jiang et al., 2013; Mohanakrishna et al., 2016). A low applied potential develops better biofilm on cathode electrode whereas high applied potential would increase the H2 evolution rate in the biocathode chamber (Mateos et al., 2018; Mohanakrishna et al., 2016). Applied potentials shall be as low as possible for practical consideration of BES. Instead of increasing the cathodic overpotential, the cathodic media can be supplied with external H2 and CO2 to reduce electricity costs. It is necessary to optimize the potentials and understand the implications better before implementation at a practical scale.

6.3.6 Effect of reactor design The shear forces produced by the different hydrodynamic conditions employed by various BES configurations may affect the structure, and activity, of microbial communities underpinning the BESs, as well as affecting the mass transfer, that is, the exchange of substrate and products between the anolyte and the anode biofilm, which could, in turn, impact on power output (Cecconet et al., 2018). H type cells are the simplest and the first-ever used reactor in the electrochemical field (Nevin et al., 2010) as well as for MES. Though simple in construction, this type of cell possesses high internal resistance due to the distance between the electrodes (Park et al., 1999). The modified form of H type cell is the cube cell reactor and flat plate reactor. In these rectors the distance between the electrode are less compared to the H cell. Apart from the these three flat reactors, concentric tubular reactors are also used (Pozo et al., 2015). In a column/up-flow-type reactor, the cathode is placed above the anode. The anode can be designed as a fixed bed to provide better stability and support to the attached microbial biofilm (Guo et al., 2010). Apart from these, other reactor models are available, which are developed by different researchers to fulfill different requirements like rotating disk reactor, cylindrical

6.4 Strategies to improve product titer

reactor, stirred tank reactor, and reverse dialysis stack reactor (Bo et al., 2014; Krieg, n.d.; Luo et al., 2014). Some of the most used reactor configurations and their performance are mentioned in Table 6.2. In BESs the overpotentials unavoidably increase with the reactor size (Rossi et al., 2019); thus, a small, compact, and stackable reactor design is preferable for scaling-up rather than a single, high-volume reactor (Greenman and Ieropoulos, 2017). Stacks with several MES modules also facilitate maintenance, since individual modules requiring maintenance can be stopped while keeping the other modules under operation. Multichamber reactors with flat (Bajracharya et al., 2016; Srikanth et al., 2018a) or tubular (Batlle-Vilanova et al., 2017a; Blasco-Go´mez et al., 2019; Pepe` Sciarria et al., 2018) structures are examples of easily scalable and stackable configurations to increase the electrode/surface ratio and reduce footprint (Flimban et al., 2019).

6.4 Strategies to improve product titer Comparison of all the published MES studies is difficult because the production rate is dependent on different parameters (electrode material, inoculum, pH, and applied potential). Statistical analysis was performed for comparison of the different parameters by collecting the data from various studies, which are relevant to each parameter (two or more values considered in the same study) for MES performance (Fig. 6.2). The MES performance (P), in terms of acetate production studied with each parameter (electrode material, pH, inoculum, and applied potential) was normalized to the maximum MES performance (Pmax) found in the same study. Fig. 6.2 represents that P/Pmax near or equal to one indicates that a value of single parameter studied resulted in a higher production rate than the other value of the same parameter in the same study. Carbon/graphite plate and carbon cloth electrodes resulted in poor performance due to less bacterial attachment on its surface. Modified materials have shown high performance (P/Pmax close to one) compared to other electrodes due to its high biocompatibility (Fig. 6.2A). In contrast, chemically treated electrodes though enriched performance are not suggested because of their increased cost. Applied potentials below 20.7 V versus standard hydrogen electrode (SHE) have shown higher P/Pmax value than the other negative potentials. From Fig. 6.2B, it is interpreted that the maximum production rate has been achieved at more negative potentials. It is due to the high number of electroautotrophic microbial species able to grow under such negative potentials. Fig. 6.2C shows that, the pH between 5 and 6 resulted in a higher rate of acetate production, which is due to high rate activity of biocatalyst and faster biocathodic reduction reaction on the cathode at near-neutral pH. Electroactive bacteria are more sensitive to the pH of catholyte, which is why pH below 5 resulted in poor MES performance (lower production rate). BES-enriched inocula from earlier MES produce, on average, slightly higher acetate production than other inocula (Fig. 6.2D). This is generally due to the higher number of electroactive microorganisms presented in the enriched inocula.

131

Table 6.2 List of different reactors used in microbial electrosynthesis from the previous studies.

Cathode

Reactor material

Cathodic volume (mL)

Product

Graphite plates

Graphite plates

Glass

2.5

Acetate

Dynamically stable anode Dynamically stable anode Titanium mesh Dynamically stable anode

Graphite stick with two felt Graphite plates

Glass

250 3 2

Acetate

3.55 mM/L/ d 10 g/L

Glass

250

Acetate

N.A

Carbon felt Graphite

Glass Glass

250 250

Acetate Acetate

3.92 g/L N.A

Polished graphite plate Active carbon

Hybrid metal plates

Glass

100 3 2

Acetate

1.23 g/L/d

Active carbon

Borosilicate made

500

Acetate

Pt plate with Ti plate current collector Graphite stick

Graphite felt and stainless steel mesh Graphite stick

PVDF plastic

500 3 2

Acetate

24.53 mg/ L/d 1.3 mM/d

Plastic

250 3 2

Carbon felt

Carbon cloth

Glass

200 3 2

Wax ester Acetate

Activated carbon cloth with graphite rod Carbon cloth coated with Pt (20%) Carbon rod

Graphite plate loaded with carbon power Mxene coated carbon felt Graphite felt

Glass

N.A

Acetate

58.2 63.2 mM/d 7.2 g/L

Plexiglass

200

Acetate

765 mg/L

Srikanth et al. (2018b) Tahir et al. (2020)

Plastic

14 3 2

Acetate

N.A

Hou et al. (2020)

Type of reactor

Anode

H type

Cube type

Rate of production

38 µM/L

References Annie Modestra et al. (2015) Bajracharya et al. (2017a,b) del Pilar Anzola Rojas et al. (2018b) Dong et al. (2018) del Pilar Anzola Rojas, Zaiat, et al. (2018a) Jiang et al. (2019) Mohanakrishna et al. (2020) Bajracharya et al. (2015) Lehtinen et al. (2017) Yu et al. (2017)

(Continued)

Table 6.2 List of different reactors used in microbial electrosynthesis from the previous studies. Continued Type of reactor Concentric tubular

Reverse dialysis stack Cylindrical chamber

Reactor material

Cathodic volume (mL)

Product

Anode

Cathode

Pt or NanoWeb RVC EPD-3D Platinum counter electrode Carbon fiber mesh

Graphite plate or NanoWeb RVC EPD-3D Cylindrical graphite felt

Glass

300

Acetate

Glass Glass

300 500

Acetate Acetate

Ti coted pt mesh

Plastic

100 3 2

Carbon rod

Graphite felt

Glass

15 20

PVDF, polyvinylidene difluoride.

Rate of production

References

1.3 mM/ cm2/d 685 g/m2/d 261 mg/L/d

Jourdin et al. (2014)

Acetate

165.79 mM/d

Li, Angelidaki, and Zhang (2018)

Acetate

550 mg/L

Song et al. (2020)

Jourdin et al. (2015) Mateos et al. (2020)

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FIGURE 6.2 Data analysis on A) Effect of surface electrodes. B) Effect of applied electric potential (V vs. SHE). C) The effect of pH. D) Impact of inoculum. The P/Pmax represents the MES performance in terms of acetate production rate with the earlier-reported parameters compared to the maximum MES performance obtained in the same study. N refers to the number of papers considered for the statistical analysis.

According to our knowledge, the currently available studies related to pH and inoculum on MES (3 or more) are not enough for statistical analysis. Further research studies with proper statistical analysis and design of experiments are required to determine the optimum pH, inoculum, and reactor design for the improvement of MES performance.

6.5 Economic evaluation The first step in a technoeconomic evaluation of the CO2 electrochemical conversion route is a basic screening of thermodynamic properties of reactions toward different base chemicals. This already gives a rough indication of feasibility and allows early

6.5 Economic evaluation

decision-making in screening technological strategies toward viable business cases. Production of formic acid, acetic acid, oxalic acid, methanol, and ethanol are thermodynamically feasible (ElMekawy et al., 2016; Irfan et al., 2019). The Gibbs free energy of reactions are translated into an operational cost for energy consumption, taking into account given electrical efficiencies and electricity price. This productspecific cost is then compared with the bulk price of base chemicals. This exercise was done for five base chemicals: three organic acids and two alcohols with given market price and variating energy consumptions per ton of product in the reaction of CO2 with water as reductant by ElMekawy et al. (2016). The alcohol pathway, with a more extensive reduction of CO2, is less economically feasible, where the market value is less than the cost of energy required (ElMekawy et al., 2016; Irfan et al., 2019). A similar economic assessment for the production of five medium chain fatty acids, namely, formic acid, acetic acid, propionic acid, methanol, and ethanol was reported in a recent study by Christodoulou et al. (2017). They observed that production costs were competitive with the market only for formic acid (0.30 d/kg) and ethanol (0.88 d/kg), accompanied by high returns of 21% and 14%, respectively. Electrochemical reaction for acetic acid production uses four times (8 e2) more electrons than formic acid (2 e2) and thus results in a higher energy demand of 874.82 kJ/mol as compared to 269 kJ/mol for formic acid. Another major factor affecting the energy consumption is the amount of water molecules produced, which tend to dilute the desired chemicals leading to energy-intensive separation processes. Production rate and energy use are the two major factors that determine the economic viability of a process (Christodoulou et al., 2017). Marshall et al. (2013) reported that a 1000 L reactor-generating acetic acid at a rate of 1 g/L cathode vol/day (1 kg acetate produced/day) at a Coulombic efficiency of 69% would require $0.35 of electricity (assuming 1.5 V, $0.05/kWh) to produce $0.6 of acetic acid (1 kg). In general, an investment in electrochemical technology becomes more attractive when targeting high energy-containing molecules, ideally having a higher market value. Considering the ratio of value and cost, oxalic acid should be the target compound, with a ratio of around 7, followed by formic acid (ratio, 3) and acetic acid (ratio, 1.5) (Elmekawy et al., 2016). Formic, acetic, and oxalic acids are currently produced via chemical synthesis routes, starting from carbon monoxide (CO) and methanol or sugar as raw material in fermentation processes. Using BES to coproduce these raw materials, like methanol in the two-step formic acid synthesis, could be a good incentive for these chemical industries to commercialize BES. The use of CO2 as raw material can be considered as a revenue stream or cost, depending on the business model. However, this incentive may become more significant for increased value of CO2 credits in future through legislation. ElMekawy et al. (2016) observed that the energy cost contributes to more than 50% of the production cost and is most significant in case of acetic acid production. Other than economic evaluation, the market size is an important factor before establishing a business case, where high throughput is required in the electrochemical conversion of CO2 to justify investments and an acceptable payback

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period. Market saturation is an important limitation in defining the production strategy. For example, oxalic acid is the most interesting base chemical from an economic point of view, however if 30 kton product is produced per year, that corresponds to oxalic acid market share of around 10%, which is around 4% for formic acid and just 1% for acetic acid. A possible strategy to overcome this issue is the differentiation in targeted end products. The two major substrates in MES is CO2 and organics, where electrical current is used for fixation of CO2 or conversion of organics. While there are several advantages of using CO2 as a substrate for bioproduction, such as unlimited availability (atmosphere, waste gas, etc.), land-independence, no external energy requirement for uptake, and limited toxicity to the microorganisms, production rate of fatty acids are much smaller than that compared to organic substrates (Desloover et al., 2012). Also since CO2 is the most oxidized state, a large number of electrons are required for product formation. For example, production of 1 mole of succinate from CO2 would require 14 electrons, while it will require only 2 electrons if produced from glycerol. Also the elongation of short-chain fatty acids such as acetate to caproate is the only way to synthesize long-chain fatty acids. Therefore, the use of more reduced substrates like glucose, acetate, butyrate, lactate, etc. can be an interesting alternative as it significantly lowers the electron requirement and thus power demand of the production process. In most cases, obtaining pure substrate as a starting point will be too expensive, considering the initial refining required. However, there is a considerable supply of waste organics with less value like glycerol- and fatty- acid-sources. Transforming glycerol to propane- 1,3-diol gives a significant value increase. It is perceived that, at least in the short span, organicsbased conversions can gain higher returns on investment. Besides, the acceptable system cost is a factor of 10 folds higher (h232 1684 per m2 installed) than the production from carbon dioxide (Desloover et al. 2012).

6.6 Future scope of work In the future, the MES process may get competence to meet market demand in supply of valuable chemicals. So far, the extracellular electron transfer mechanism is not clearly understood. Understanding the fundamentals of electron transfer could improve the process performance (Marshall et al., 2013). Also, the research focus could be on gene modification of acetogens to divert pathways for the production of long-chain acids. There is a necessity to develop biocompatible cathodes at cheap cost to achieve high-current density (for H2 evolution). Further efforts on electrode material improvement are needed to enhance the rate of production by selecting novel and low-cost materials. The MES process could be enhanced using three-dimensional electrodes, novel reactor configuration, and production of long-chain fatty acids at a higher rate (Jiang and Zeng, 2018). To scale up the MES reactor, the use of stacked MES reactors could be more efficient than merely increasing the size of a single reactor. A large single reactor

References

would inevitably increase potential losses due to the increased electrode distance (Liang et al., 2018). The optimization of operating parameters would reduce the overall cost of the BESs to make them economical (Aryal et al., 2017a). Incorporating renewable power supply to MES helps to reduce production cost and can bring the process more close to practical level application. MES can be applied for retrofitting existing power plants by storing renewable energy during peak production, or by producing organic substrates as organic sources for biological wastewater treatment (e.g., denitrification). Electrosynthesis products such as H2, CH4, CO, and short-chain fatty acids can be further upgraded to longer-chain carboxylates, biofuels, bioplastics, polysaccharides, and protein with multistep bioconversions. MES can also be seen as part of a biorefinery aimed at cleaning wastewater (at the anode) while generating upgraded biogas (at the cathode), and recycling the resulting CO2 to products for other purposes. Similarly, MES can provide reducing equivalents for syngas fermentation. Biofuels produced from CO2 can be used within the company as alternative fuel for boilers, CHP plants, and transportation.

6.7 Conclusion MES is a flexible technology for production of various marketable products from CO2. This book chapter summarizes the parameters affecting product yield in MES. Currently, MES technology is producing low-value organic acids at a higher rate, but it is required to target high-value long-chain products (nC2, nC4, and nC6) with a high rate. Maximum production rate and purity, with minimum costs and footprint, are required for MES scale-up. MES can be applied to retrofit existing treatment plants to comply with the future strict legislation on emissions. In-line extraction of marketable product separation will also result in higher production rates. A combination with wastewater treatment, use of renewable energy, and integration of energy harvesters can make scaled-up MES.

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CHAPTER

Low carbon fuels and electro-biocommodities

7

Bahaa Hemdan1,2, S. Bhuvanesh3 and Surajbhan Sevda4 1

Department of Biosciences and Bioengineering, Indian Institute of Technology Guwahati, Guwahati, India 2 Water Pollution Research Department, Environmental Research Division, National Research Centre, Giza, Egypt 3 Director’s Research Cell, CSIR-National Environmental Engineering Research Institute, Nagpur, India 4 Department of Biotechnology, National Institute of Technology Warangal, Warangal, India

7.1 Introduction Expedited economic and technological evolutions around the world have accelerated the production of tremendous amounts of wastes in three different states, including solid, liquid, and gaseous form. The wastes produced pollute the soil, water, and the atmosphere with dangerous and destructive chemicals. The humongous generation of wastewater and its discharge in water bodies leads to polluted water sources. Industrial carbon dioxide emission is also one of the massiveconcerns of waste, which could create a significant threat to the ecosystem and lead to global warming and climate change (Srikanth et al., 2016). Unregulated dumping of toxic wastes causes heavy metal pollution in water and soil (Vongdala et al., 2019). Relentless burning of solid wastes produces many dangerous gases such as CO, CO2, SO, NO, and other carcinogenic emissions altering the environment (Little et al., 2018). The maladministration of Municipal Solid Waste (MSW) is indeed responsible for the substantial and multifaceted environmental and communal consequences that do not enable adjustments in environmental sustainability (Ferronato and Torretta, 2019). Bioelectrochemical systems (BESs) are groundbreaking and innovative platforms for bioengineering that use electrochemically active bacteria (EAB) to convert stored chemical energy into bioelectricity, hydrogen, or biochemical products such as H2O2, acetate, and ethanol (Martinez and Alvarez, 2018; Naha et al., 2020; Sevda et al., 2020). In BES, intentional growth of EAB can be used as a biocatalyst for bioelectricity generation, wastewater handling, and H2 formation. This technology provides a responsible approach for extensive waste management, electricity, and resource restoration in comparison to conventional approaches (Chatterjee et al., 2019). The implementation of BES in the sectors Delivering Low-Carbon Biofuels with Bioproduct Recovery. DOI: https://doi.org/10.1016/B978-0-12-821841-9.00004-9 © 2021 Elsevier Inc. All rights reserved.

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of waste management can be used in the production of biocommodities, degradation of the broad spectrum of biological wastes, desalination of brackish water, decolourization of textile wastes and reduction in carbon dioxide emissions (Naha et al., 2020). This multidisciplinary system relies on the usage of EAB that can transfer formed electrons to the electrodes. This research area operates on the interface between biochemical engineering and electrochemistry, although specific studies such as electrode development, novel ion-exchange membrane research, and suitable EAB microbes search are also important for the system completion (Sevda et al., 2020). BESs reapplied not only for generating electricity but also for producing electro-bio fuel products (Sharma et al., 2014). It depends on where the physiological electrons are harvested from the electronic donor or delivered to the electrode receiver (such as the steel electrode) via electrochemical reactions (Rosenbaum and Franks, 2014). Some bacterial strains, like the Escherichia coli, produce energy through the oxidation of materials by suppressing the redox kinetics to produce an electrical circuit (Thirty-fourth Annual Meeting of the Society of American Bacteriologists, 1933). The understandings in the bioelectrochemistry field can indeed be traced back to the first half of the twenties century with the outcomes by Potter (2013).

7.2 Working mechanism of bioelectrochemical systems In BES, oxidation of wastewater containing organic materials is carried out in the anodic chamber to harvest electrons (e2) and protons (H1). The protons are transferred to the cathode where they are employed in redox reactions using EAB (Wang and Ren, 2013; Nesakumar et al., 2019). The EAB microbes for transferring electrons involve many cell surface extracellular electron transfer (EET) pathways including (1) direct connection: the connection between the external membranes of the electroactive bacterial cell and the electrode surface as an external electron receiver; (2) indirect contact: electron transferable through an electron quill handling such as water-soluble electron carrier particles behaving as electrochemical negotiators, having to block electrons from the surface of electrodes and bacteria; and (3) using penetrable pills or pills as constructions (Rabaey et al., 2010). As shown in Fig. 7.1, BES is further classified into different types based on its construction and application such as microbial fuel cell (MFC), microbial solar cell (MSC), microbial electrolysis cell (MEC), microbial desalination cell (MDC), and microbial electrosynthesis (MES). In MFC, the electrons and protons are transformed instantly to produce bioelectricity (Slate et al., 2019; Rabaey et al., 2005; Sevda and Abu-Reesh, 2017). On the whole, MEC performs under an anaerobic circumstance that facilitates the formation of H2, alcohol, and other valued chemicals in the cathodic chamber. In the MEC, oxygenic photosynthetic and pigmented organisms (cyanobacteria) or higher plants (autotrophs) are degraded

7.2 Working mechanism of bioelectrochemical systems

FIGURE 7.1 Diagrammatic description of the bioelectrochemical sciences, the interdependency with the bioelectrochemical systems.

by the EAB to produce electrocommodities by application of an electric current (Wang et al., 2015). MDC are used in desalination coupled with hydrogen recovery (Zhen et al., 2017; Zhang and Angelidaki, 2014). The generated electrons in the cathode chamber can also be utilized to generate energy and value-added goods (Karthikeyan et al., 2019). In BES reactors, generated bioelectricity or commodities are from sustainable and carbon-rich wastes, generating minor greenhouse gases (GHG) discharges in comparison with traditionally applied approaches. In BES processes, the versatility of the system or the implemented small voltages produced in situ lets the biocatalyst to develop the individual item in overcoming the thermodynamic vitality boundaries. Wastes with negative value is sometimes used as a substratum to build vitality or fill in BES because the design provides the critical electrons required for the reactions (Abdel-Shafy and Mansour, 2018). Various MSW and wastewaters collected from different sources comprising diverse forms of organic components are perceived as primary substrates for the anode chamber in BES. Nonetheless, the characteristics and structure of waste determines its proficiency in BES in terms of bioelectricity generation. MSW contains low water and high concentrations of solids/natural materials that are not well treated. Therefore, wastewater may indeed be a possible carbon hotspot for MFC due to the possibility of turning renewable and priceless energy from harmful waste (Kumar et al., 2019). There are exchange rates between such measurement techniques and should be monitored through appropriate and flexible processes (Fig. 7.2) (Raadal et al., 2011).

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FIGURE 7.2 Bioenergy associated dynamic system relationships between civilization, energy, and the environment. Strategies to disorganized food and fuel production emerging from inadequate land-use management are instances of daily business.

Many bioenergy solutions can offset global changes in the atmosphere if they are replaced by conventional non-renewable energy sources, such as the use of burning fossil fuels and low bioenergy emissions. Generally, applied novel methods for treating different types of waste streams incorporate thermosynthetic processing and biohelped techniques. The innovations for solid waste replacement, just as liquid and vapors waste handling, are energy escalated and antagonistically influence the ecological respectability of science (Srikanth et al., 2016). The more significant part of these procedures discharge CO2 into the earth, causing expanded global temperatures.

7.3 Application of microbial electrochemical technologies in wastewater treatment The use of microbial electrochemical technology (MET) depends on the occurrence of the metabolic function of EAB for the utilization of solid-state electrodes for various oxidizing types of wastewater (Chiranjeevi and Patil, 2020). Several types of wastewater can be generated from numerous industrialized resources that involve organic compounds that contribute significant support for bioenergy

7.3 Application of microbial electrochemical technologies

creation. For instance, oxidative wastewater strategies under aerobic conditions are impractical due to regular power requirements for aeration and sludge handling (Ram´ırez-Vargas et al., 2018). At present, MFC innovation presents a proper option for energy-positive wastewater treatment with bioelectricity generation and asset recuperation through bioelectrochemical remediation. The additional benefit of utilizing the MFC approach for beginning wastewater treatment is that few bio-based procedures including expulsion of biochemical and concoction oxygen request, nitrification, denitrification, sulfate removal, and removal of substantial metals can be performed in the bioreactor (Kumar et al., 2019). A significant energy capacity in the form of biodegradable organic matter is stored in wastewaters. Furthermore, MET has demonstrated the capacity for waste management as their chemical oxygen demand (COD) removal levels could reach up to 90%, with coulombic efficiencies up to near 80% (Singh et al., 2019a). It is reported that 1 kg COD can be transformed to 4.16 kWh of power in an anaerobic bioreactor (Cao et al., 2019). The BES technology is considered as an emerging technique for harvesting bioelectricity from wastewater (Lovley, 2006; Corbella et al., 2015). However, the MET technology having few drawbacks that includes; the cost of configuration, expensive electrode materials, and lower power density (Kabutey et al., 2019). Moreover, for the wastewater treatment, METs are an excellent choice since they can swiftly oxidize the degradable organic matter using biocatalyst (Rabaey et al., 2007). Additionally, METs are versatile platforms that really can handle wastewater at a rate of 7.1 kg COD/m3/d by oxidation and reduction procedures, and implement the interest of saving expenses such as aeration and sludge management compared to the conventional methods (Maiti et al., 2019). The MET system also provides a new method for simultaneous wastewater treatment and biocommodities production in the cathodic chamber (Rosenbaum and Franks, 2014). The most studied usage of MET is bioelectricity generation with wastewater treatment using MFC. The EAB microorganisms are used in the anodic chamber and anode electrode works as an electron acceptor to degrade the organic composites biologically using oxidizing agents. The generated electrons retransferred via an external electrical circuit to the cathode resistance (Slate et al., 2019). The generated bioelectricity can be recovered from the outer circuit, as the chemical oxidation reduction reactions are thermodynamically desirable (Santoro et al., 2017). The MFC configuration is generally classified as a dual or single-chamber constructed technological system (Fig. 7.3). The bioelectricity generated can be retrieved from the outer circuit, as in a dual-chamber configuration, both anode and cathode is immersed in electrolytes that promote both oxidative or reductive reactions and the flow of protons by an ion-exchange membrane and the electrons along an external circuit. In the single chamber, anode and cathode electrodes are submerged on a similar electrolyte, or the cathode can be conferred to the air. In this setup, the released protons relocate to the cathode by a polymer electrolyte film and the electrons through an external circuit (Gude, 2016).

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FIGURE 7.3 Descriptions of microbial electrochemical technology installations for treating wastewater, spontaneous reactions: the standard microbial fuel cell with double-chamber.

7.4 Electro-biocommodities and value-added biochemical’s production It has been long identified that BESs can have utilization other than wastewater treatment and electrical power formulation (Hamelers et al., 2010). The term “microbial electrosynthesis” was begotten in 2010 for the decrease in the levels of carbon dioxide emissions to the generation of multicarbon commodities with electrons transmitted from a terminal as the electron contributor (Kumar et al., 2019; Pant et al., 2012). The MECs are a generally unique concept compared with MFCs. The biological reduction of CO2 into multicarbon mixtures at the cathode utilizing chemolithoautotrophs is a rising utilization of MES (Bajracharya et al., 2015). The capturing and employment of CO2 has recently become an appealing strategy to reduce ever-increasing carbon patterns and greenhouse gaseous on earth and supplement the bioenergy/renewable energy (biofuel) field with formed electro biochemical commodities (Owusu and Asumadu-Sarkodie, 2016). The extra generated voltage from BESs techniques is used to for CO2 reduction to valuable products with only supply the remaining external voltage supply (Chiranjeevi and Patil, 2020). The first MES study reported in 2010 investigated that electron conversion potential of Sporomusa ovata, prompting another use of BES in the union of elements of carbon dioxide (Nevin et al., 2010). The MES incorporates a power-driven combination of capacities and synthetic concoctions, utilizing microorganisms as a biocatalyst (Pant et al., 2012; Christgen et al., 2020). The organisms that connect on the cathode surface are considered modest self-recovering biocatalyst and robust access to deliver oxide-

7.4 Electro-biocommodities and value-added biochemical’s production

decrease responses adequately when contrasted with the utilization of other costly synthesis purposes. A wide assortment of significant items can be delivered to hydrogen, ethanol, methane, acetic acid derivation, butanol, and H2O2. Fig. 7.4 shows an outline of the MES system working principal for treating wastewater with biocommodities production.

7.4.1 Biohydrogen production Hydrogen is used as a power transporter and has already been discovered to be an appealing source of renewable energy. It is commonly used in various industrial advancements as a primary reference point for diesel/chemical, owing to the lack of hydrogen levels in the environment. Moreover, the biohydrogen generation mechanism also has certain inefficiencies, such as thermodynamics and low output of microbial metabolic molecules that arise during the fermentation phase (Saratale et al., 2015; Singh et al., 2019b). The BES is considered a feasible alternative strategy to objectively enhance biohydrogen generation at low electrical intensity (0.2 0.8 V) compared to the traditional water electrolysis, which requires high electrical supply with high input of 23 1.8 V (Chatterjee et al., 2019; Yu et al., 2018). Certain species of bacteria with electrical impulses can degrade organic molecules, especially in the anode chamber with acetate in bioelectric hydrogen, and produce carbon dioxide, protons, and electrons. Whenever the variable that determines is regulated, the electrons pass towards the cathode electrode and thus associate with available protons to biologically generate biohydrogen in the cathodic chamber (Logan and Rabaey, 2013). While the yield of generated biohydrogen is more prevalent in this process at multiple times (11 mol H2/mol glucose) compared to the dark fermentation process (Liu et al., 2010; Wheeldon et al., 2017). The use of platinum (Pt) and nickel (Ni) are enhanced with the biohydrogen generation rates compared to the metal electrodes (Buitro´n et al., 2018).

7.4.2 Biomethane production Biological electrosynthesis of biomethane that employs treated outlets from anaerobic bioreactor has many advantages, and along with the technique could be conducted at colder temperatures; it would be less sludge-forming and demands no air circulation expenses and higher output of biomethane, which enables approaches more environmentally sustainable and affordable (Borole, 2016). Both the anode and the biocathode developed by a biofilm of Methanobacterium palustre were conceived in the MEC bioreactor to generate biomethane. An individual compartment designed to create methane at a rate of 0.75 6 0.12/L-MEC/d with the electrode manufactured from plain graphite (Moreno et al., 2016). After 16 days of operation with dark fermentation waste products, biofilm formation in the cathode turned out to be a revolutionary and successful system with outstanding material transfer performance (Liu et al., 2019). The MES is an ideal

149

FIGURE 7.4 Illustration of a microbial electrosynthesis system for the treatment of wastewater and generation of biocommodities in a single reactor.

7.4 Electro-biocommodities and value-added biochemical’s production

opportunity to convert carbon dioxide into biomethane. In principle, it is well understood that directed biochemical reactions generate methane from carbon dioxide by sending the electron explicitly, or often use hydrogen or acetate or indirectly coordinate. Firstly, CO2 reduction has also been identified as a combination of methane, acetate, hydrogen, and shape using a dynamically mixed microbial culture in a beverage factory with a capacity of 2590 mV (vs SHE) where CO2 was the only chemical compound considered (Ammam et al., 2016). Recently, Liu et al. (2016) observed that BES utilized organic material under colder temperatures in anaerobic digestion (AD), which may be favorable for improved biomethane production. Incorporated BESs-AD system, with a cathode propensity of 2900 mV (Ag/AgCl), which produced decreased methane (31 mg CH4-COD/g VSS) and was reported to be better than in the 10 C biogas reactor. Additionally, Feng and Song (2016) examined the anode graphite fiber fabric that has been updated by different strategies to enhance the performance of biomethane emissions and to illustrate its outcomes. The generation of improved sustainable biomethane can be researched by selecting EAB during the MES process by introducing the CO2 provided by various biochemical reactions (Liu et al., 2016). Other important factors such as reactor design type, electrode design, microbial population expanding at the cathode, and other physical, operational circumstances also influence the appearance of biomethane generation (Park et al., 2018). Biomethane has the additional benefit that it could be stored or transported feasibly. Compression, pipe transport, and storage include mature technological innovations and could be incorporated quickly into existing facilities (Ishii et al., 2019). Biomethane-generating MECs are recommended as a fuel-friendly effluent cleaning stage for composting effluents, most likely with low levels of production of sludge and without aerobic treatment charges (Nakasugi et al., 2017). Biomethane generation by reduction of CO2 at a MEC biocathode with a culture media of M. palustre by electromethanogenesis has already been conducted efficiently (Geppert et al., 2019).

7.4.3 Bioethanol production Bioethanol currently contributes to about 80% of the total available renewable liquid fuel production in the world. The combination of BESs with the waste streams produced in bioethanol plants provides significant improvements for the recycling of precious and valuable commodities since most functional bioethanol factories have extremely low biomass to fuel transformation (Buˇsi´c et al., 2018). A possible strategy for transforming liquid biomass and agricultural waste into bioethanol is bioacetate reduction with hydrogen. The recent study reported that conversion of acetate to bioethanol using methyl viologen (MV) as facilitator was occurred in the cathodic chamber of BES reactor (ter Heijne et al., 2019). The generation of bioethanol had coulombic efficiency (CE) of 49%. Moreover, MV inhibited side emotional responses such as methanogenesis and improved bioethanol generation, but was gradually exhausted owing to irreversible cathode meaningful contraction

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and found that high butyrate harvests (an embarrassing end outcome) and methanogen in its absence (Chen et al., 2016; Contreras-Da´vila et al., 2020). In the MES, use of acetate as a substrate for bioethanol generation in the cathodic chamber with a capacity cell voltage range of 20.28 V. Numerous reports have already shown that various components of gasification (CO, CO2, and H2) have been transformed into multicarbon mixtures by applying acetogenins, but the ethanol has been formulated as an inconsequential end-outputs (Flexer and Jourdin, 2020). Microbial species that reduce the sulfur component of the cathode were formed to develop bioethanol and many other oily compounds to convert acetic acid and butter acid by transferring the vital energy at 20.65 V (Ruffine et al., 2018). The generation of other alcohols, such as methanol (0.8 mM), isopropyl alcohol (0.4 mM), ethanol (0.2 mM), butanol (0.6 mM), acetone (0.2 mM), and a small proportion of propionic acid, caproic acids were also reported in a BES reactor (Bajracharya et al., 2016b). Electrolysis by genetically modified Ralstonia eutropha succeeded in the production of biohydrogen and the further reduction of CO2 to form liquid carbon fuels. This modified bacterial species could withstand electrical charge and redirected the usual polyhydroxybutyrate formation path and generated isopropanol (up to 216 mg/L) (Pant et al., 2019). Similarly, the formulation of bioethanol in BESs can be stimulated by promoting or enhancing the microbial activity of the biocathode electrode by methods of a nonmediated consumption of acetate (Bajracharya et al., 2017). The biologically induced cathode could produce water and electricity in the presence of oxygen as in the situation of MFCs, and/or MECs generate biohydrogen under anaerobic conditions. Eqs. (7.1) and (7.2) illustrate oxidation of monosaccharide (glucose) into the carbon dioxide in the anodic chamber and water formation in the cathodic chamber. Anode:C6 H12 O6 1 6H2 O-6CO2 1 24H1 1 24e2 E0 5 0:014V

(7.1)

Cathode:24H1 1 24e2 1 6O2 -12H2 OE0 5 1:2V

(7.2)

7.4.4 Acetate production Acetate is considered the most critical extracellular middle biochemical product throughout biological generation (Hubenova et al., 2015). It also is the leading CO2 conversion output documented, but numerous recent research has revealed that it is conceivable to concurrently reduce CO2 to the blend of acetate, butyrate, and ethanol using mixed bioactive microbial species (Bajracharya et al., 2015). Further, the valuable fuels and electro biofuel goods could be performed owing to the biological reduction of acetate (Sharma et al., 2013). Nevin et al. (2010) investigated that when Sporomusa ovate inoculated in the cathodic chamber at a cathode voltage of 20.4 V, acetate production occurred in the MEC reactor. Cathode electrons convert CO2 and contribute to producing acetate, 2-isobutyrate, or format. Comparable reports were also conducted utilizing miscellaneous microbial species where the cathode charge was lower than or equal to 20.59 V (Kracke et al., 2019). Anode-containing sulfide using Desulfobulbus propionicus

7.4 Electro-biocommodities and value-added biochemical’s production

was productively discovered and conducted as an electron provider, most of which transformed over CO2 into the acetic acid generation in the cathode using S. ovata as a biocatalyst in biological electrosynthesis (Gong et al., 2013). Additionally, the high rate of methane and acetic acid derivation generation from CO2 utilizing biotic cathode by blended culture (Aryal et al., 2017). It is reported that only certain classes of acetogenic microbes, that is, Clostridium aceticum, Sporomusa sphaeroides, Moorella thermoacetica, and Clostridium ljungdahlii, seemed to use electrical current to generate organic compounds such as 2-oxobutyrate acetate and formate (Pre´voteau et al., 2020). Carbon dioxide was used as a substrate in a comparative study with the progressive typical microbial biofilm and reduced to the derivation of acetic acid (1.5 g/L) (Marshall et al., 2012; Batlle-Vilanova et al., 2017).

7.4.5 Hydrogen peroxide production The generation of hydrogen peroxide (H2O2), a significant industrial commodity in the BES reactor. The production of H2O2 was reported based on bioelectrochemical oxidation of sewage organic materials combined with cathodic CO2 conversion to H2O2 (Zou and He, 2018). This acetate-based system produced 1.9 kg H2O2/m3/d of H2O2 during external imposed cell voltage of 500 mV (Chen et al., 2014). The generation of H2O2 was reported using a microbial reversed electrolysis cell reactor with an imposed external voltage of 2485 mV (vs Ag/AgCl). The average cumulative concentration of H2O2 was 778 6 11 mg/L (Ogawa et al., 2018; Bajracharya et al., 2016a). In addition to the obvious BESs applications as a platform for bioremediation and pharmaceutical product synthesis, they may also be used as imaging techniques.

7.4.6 Other value-added biochemical production The BES reactor generates high-value commodities such as fatty acids and alcohols. Butyric acid is well respected as an advantageous and possible molecule that is frequently used in various industries such as pharmaceutical and textile management as a forerunner to biofuel (butanol) (Bian et al., 2020; Zhang and Tremblay, 2018). The butyric acid generation in BES reported with using with Clostridium tyrobutyricum BAS7 at the cathode compartment at 2400 mV cell potential using normal red as an aqueous solution showed significant butyrate output of 8.8 g/L (Vassilev et al., 2019). MV, present at the cathode as an electron mediator, was able to reduce acetate for the production of bioethanol. Besides, BESs not only manufacture syngas but also produce high-value organic molecules, including fatty acids and alcohols. Such organic products are usually used as the principal precursor for the manufacture of certain polymeric materials, plasticizers, and biodiesel outcomes and as lubricants in various industry quarters (Fu et al., 2019). Ammonium hydroxide was produced in BES reactor at a rate of 0.09 mM/L/h through CO2 conversion and with 64.8% of CE (Leung et al., 2020).

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7.5 Recent progress for electro-biocommodities generation in a bioelectrochemical system The use of inexpensive and premium materials for the construction of electrode and membrane, such as ceramic membranes or nonplatinum catalysts, also empowers us to achieve a viable power generation device and biocommodities in the BES reactors (Palanisamy et al., 2019). Despite this, the performance of MFCs, by necessity, are enhanced before they can be developed for commercial purposes for wastewater treatment and valuable low-cost biocommodities (Herna´ndez-Ferna´ndez et al., 2015). The EAB microbes used for the production of green power, bioremediation, and other commercial applications in BES are promising at lab-scale experiments, but more research work needs to be done for their commercial economical processes. Table 7.1 show the various biocommodities generation in various BES reactors. Moreover, the EET mechanism in BES is restrained to element-reducing significantly different microbes, particularly in the Geobacter spp., and Shewanella spp., but certain aspects of other EET-involved ingredients require to be discovered (Choudhury et al., 2017). Since these different microbes already dedicate a significant amount to the MFCs, it also is essential to identify other exoelectrogens engaged and reached for elevating the collaboration of technological approaches supported by MFC (Li et al., 2018). A more comprehensive perspective into the biofilm features and genetic accomplishment of organisms can hence admit up original opportunities for promoting an MFC’s competence (Angelaalincy et al., 2018). The biogenetic engineering of such electroactive organism for increasing production of the deep layer of biofilm and enhancement of additional electron transfer pathways that lead to an escalation in the existing capacity of an MFC and improved extraction of COD in sewage processing facilities equipped with MFC strategy (Xin et al., 2020; Babauta and Beyenal, 2015). Extracellular sugars generated by microbes are associated with the establishment of biofilm and the exchange of electrons. Since the cell surface modification, media optimization techniques, and confirmation of operating conditions are distinguished as fundamental aptitudes of electrochemical enforcement, the biotechnology of the primary genetic modifications critical in the development of electrochemically active biofilm is considered to be the expected outlook (Angelaalincy et al., 2018). Broad assortments of photosynthesizing and anaerobic microorganisms have been engaged as electron donators and acceptors in MFCs. The most predominant species, which could be utilized in MFC, include Pseudomonas fluorescens (Friman et al., 2013), Chlorella vulgaris (Commault et al., 2017), Phormidium sp. (Bradley et al., 2012), Saccharomyces cerevisiae (Permana et al., 2015), Geobacter metallireducens (Kanjilal et al., 2017; Poddar and Khurana, 2011), Thiobacillus ferrooxidans (Zhang et al., 2019), Desulfovibrio desulfuricans (Kang et al., 2014), Leptothrix discophora (Rhoads et al., 2005), Scenedesmus armatus (Angelaalincy et al., 2017), and Rhodospirillum rubrum (Bensaid et al., 2015). A few of these microbes are genetically modified to achieve accelerated outcomes in the case of bioenergy generation

Table 7.1 Biocommodities production in various reactors. S. No.

Substrate

Biocommodities

Flat plate reactor bioelectrochemical systems Tubular bioelectrochemical systems (2400 mV by a DC power supply)

Wet biomass waste

Steinbusch et al. (2010) Jabeen and Farooq (2016)

3

H type microbial electrosynthesis system

4

Two-chambered H-cell reactors microbial electrosynthesis system (The biocathode was poised at a potential of 21.0 V)

Average acetate production rates were around 135 mg/L/d (Pvol) and 10 g/m2/d (PESA)

Rojas et al. (2018)

5

Five identical microbial electrosynthesis system

Hydrogen-containing syngas mixture (H2/CO2 (80/20, v/v)) in cathodic chamber; artificial wastewater in anodic chamber water containing per liter: 4 g CH3COONa, 4.09 g Na2HPO4, 2.54 g NaH2PO4, 0.31 g NH4Cl, 0.13 g KCl 10 ml vitamin solution and 10 mL mineral solution (pH about 6.9) Mineral medium containing per liter: KH2PO4 (0.33 g); K2HPO4 (0.45 g); NH4Cl (1 g); KCl (0.1 g); NaCl (0.8 g); MgSO4
7H2O (0.2 g); vitamin solution DSMZ 141 (10 mL), and trace element solution DSMZ 141 (10 mL). Sodium bicarbonate (0.05 M) Phosphate buffer solution was used in the anodic chamber and Lauryl tryptose broth was used as catholyte, CO2 content from biogas

Bioethanol (13.5 6 0.7 mM) 1 H2 (0.0035 Nm H2/m2/d) Percent cathode recovery of organic products of Clostridium ljungdahlii (% ethanoic acid 3.008; % ethyl butyrate 23.463; % hexanoic acid 72.225; % heptanoic acid 94.918; % ethanol 7.843; % hexanol 96.963; % heptanol 18.925) Acetate production values were 4.80 6 0.91, 9.85 6 0.37, 12.26 6 0.31, and 16.74 6 1.07, 12.26 6 0.31, 9.85 6 0.37 and 4.80 6 0.91 mM, respectively, with the applied voltages of 0.8, 1.0, 1.2, and 1400, 1200, 1000 and 800 mV

Acetate (52.4 mM/m2/d), isobutyrate (36.2 mM/m2/d), propionate (41.6 mM/m2/d), 2-piperidinone (26.7 mM/m2/d)

Das and Ghangrekar (2018)

1 2

Reactor type

Note: PESA, projected electrode surface area.

Sterile basal medium with fructose (80 mL) in both anodic and cathodic chamber. H2/CO2 purged in the cathode electrode

References 3

Xiang et al. (2017)

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CHAPTER 7 Low carbon fuels and electro-biocommodities

and renewable and sustainable formation compared to desolate bacterial species. Also, documents related to the use of algae species in MFC have been constantly narrowed with those on bacterial species (Nandy et al., 2016; Jones and Buie, 2019). Given the role of extracellular polymeric substance (EPS) in EET of microorganisms and its relevance in biofilm formation on electrodes, the prospect of engineering the biofilm for its enhanced adhesion and EET is just the future of the MFCs. Shewanella oneidensis MR-1, a facultative anaerobe, is capable of reducing Mn (IV) and Fe (III) oxides and can produce current in MFCs. Considering the key function of EPS in the EET of microorganisms and their relevancy in electrode biofilm accumulation, the possibility of engineered organisms used to form the biofilm for its expanded conductivity and EET is only the modern world of MFCs. S. oneidensis MR-1, an essential anaerobic, can reduce the oxides of Mn (IV) and Fe (III) and of producing the voltage in the MFC device. The system occupied by EET in S. oneidensis MR-1 includes OmcA and MtrC (outer layer-OM), and c-cytes became indirect electron transfer to substantial metal oxides and anodes of MFCs; meanwhile, a Shewanella class, S. loihica PV-4 A different process is proposed to produce an electric potential (Newton et al., 2009; Oram and Jeuken, 2017). Another investigation, including cell surface polysaccharides of S. oneidensis MR-1 showed that the impact of these polysaccharides on the phone grip to graphite anodes as well as the flowage in MFCs, as the electrically nonconductive capsular polysaccharides can meddle with the contact of cytochromes of organisms to anodes and direct EET employing them. Therefore, cell surface building was prospected as an essential plan to produce higher power in the bacterial MFC system (Kouzuma et al., 2010; Cao et al., 2017).

7.6 Conclusion The production of electro-biocommodities is the most promising application of the BES reactors. In these reactors, simultaneous wastewater treatment with the generation of electro-biocommodities takes place in a single reactor system. Biohythane production using BES provides an option to control the H2/CH4 ratio in the final biohythane composition. The recent developments in the biocathodes of BES provide a new platform for generation of high-value electro-biocommodities in these reactors. Suitable reactor design, process operation control, and generation of lowcost electrodes and membrane materials are the main factors that limit the scaling up of these processes and are a strong requirement for its commercialization.

Acknowledgment The authors would like to express gratitude to Research Seed Money Grant provided from National Institute of Technology Warangal, Warangal, India for this work.

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Vongdala, N., Tran, H.D., Xuan, T.D., Teschke, R., Khanh, T.D., 2019. Heavy metal accumulation in water, soil, and plants of municipal solid waste landfill in Vientiane, Laos. Int. J. Environ. Res. Public. Health 16, 22. Wang, H., Ren, Z.J., 2013. A comprehensive review of microbial electrochemical systems as a platform technology. Biotechnol. Adv. [Internet] 31 (8), 1796 1807. Available from: https://doi.org/10.1016/j.biotechadv.2013.10.001. Wang, H., Park, J. Do, Ren, Z.J., 2015. Practical energy harvesting for microbial fuel cells: a review. Environ. Sci. Technol. 49 (6), 3267 3277. Available from: https://doi.org/ 10.1021/es5047765. Wheeldon, I., Christopher, P., Blanch, H., 2017. Integration of heterogeneous and biochemical catalysis for production of fuels and chemicals from biomass. Curr. Opin. Biotechnol. 45, 127 135. Xiang, Y., Liu, G., Zhang, R., Lu, Y., Luo, H., 2017. High-efficient acetate production from carbon dioxide using a bioanode microbial electrosynthesis system with bipolar membrane. Bioresour. Technol. 233, 227 235. Xin, S., Shen, J., Liu, G., Chen, Q., Xiao, Z., Zhang, G., et al., 2020. Electricity generation and microbial community of single-chamber microbial fuel cells in response to Cu2O nanoparticles/reduced graphene oxide as cathode catalyst. Chem. Eng. J. 380, 122446. Available from: https://doi.org/10.1016/j.cej.2019.122446. Yu, S.S., Cheng, L., Chen, J.J., Li, W.W., Zhao, F., Wang, W.L., et al., 2018. Framework of cytochrome/vitamin B2 linker/graphene for robust microbial electricity generation. ACS Appl. Mater. Interfaces 10, 35090 35098. Zhang, Y., Angelidaki, I., 2014. Microbial electrolysis cells turning to be versatile technology: recent advances and future challenges. Water Res. 56, 11 25. Available from: https://doi.org/10.1016/j.watres.2014.02.031. Zhang, T., Tremblay. P.L., 2018. Possible industrial applications for microbial electrosynthesis from carbon dioxide. In: Biomass, Biofuels, Biochemicals: Microbial Electrochemical Technology: Sustainable Platform for Fuels, Chemicals and Remediation. Available from: https://doi.org/10.1016/B978-0-444-64052-9.00032-7. Zhang, J., Ren, L., Zhang, D., Li, J., Peng, S., Han, X., et al., 2019. Reduction of NO to N2 in an autotrophic up-flow bioreactor with sponge iron bed based Fe(II)EDTA complexation. Fuel. 254, 115631. Available from: https://doi.org/10.1016/j.fuel.2019.115631. Zhen, G., Lu, X., Kumar, G., Bakonyi, P., Xu, K., Zhao, Y., 2017. Microbial electrolysis cell platform for simultaneous waste biorefinery and clean electrofuels generation: current situation, challenges and future perspectives. Prog. Energy Combust. Sci. 63, 119 145. Available from: https://doi.org/10.1016/j.pecs.2017.07.003. Zou, S., He, Z., 2018. Efficiently “pumping out” value-added resources from wastewater by bioelectrochemical systems: a review from energy perspectives. Water Res. 131, 62 73.

CHAPTER

Potential of high energy compounds: Biohythane production

8

Surajbhan Sevda1, Vijay Kumar Garlapati2, Swati Sharma2 and T.R. Sreekrishnan3 1

Department of Biotechnology, National Institute of Technology Warangal, Warangal, India 2 Department of Biotechnology and Bioinformatics, Jaypee University of Information Technology, Waknaghat, India 3 Department of Biochemical Engineering and Biotechnology, Indian Institute of Technology Delhi, New Delhi, India

8.1 Introduction Pollution levels are increasing day by day in all parts of the world. With current conditions, most of the fuel requirements are fulfilled by using nonrenewable fossil fuels. Renewable fuels have emerged as an alternative fuel compared to fossil fuels. New research is required to find the alternative fuels that emit less pollution in the environment, which are renewable in nature and also economically feasible compared to the fossil fuels. Methane is produced naturally in the anaerobic digestion of wastewater rich in organic content by methanogenic microorganisms. Methane is used widely for energy applications, and it has a calorific value of 55 kJ/g (Mishra et al., 2017). Various kinds of anaerobic digestion reactors are developed for higher efficiency methane generation and treatment of domestic and industrial wastewater in these reactors. In the current scenario, CH4 gas is used commercially in transport as compressed natural gas, household heating applications as biogas, and in the chemical industry for fertilizer generation and heating purposes. Biohydrogen is a cleaner fuel compared to methane, having a higher calorific value of 143 kJ/g (Mishra et al., 2017). Biohydrogen-based fuels are still costly compared with the current fossil fuel based economies. Biohythane is defined as mixing CH4 (70% 90%) and H2 (10% 30%) gases, and it increases the burning speed, flammability range, and combustion efficiency while reducing the total carbon emissions (Bolzonella et al., 2018). The biohythane-based energy economy is clean and sustainable compared to the current fossil fuel based economy. The combination of H2 and CH4 will reduce the overall cost and it will also increase the net calorific value, and it can be used more widely compared to CH4 and H2 alone.

Delivering Low-Carbon Biofuels with Bioproduct Recovery. DOI: https://doi.org/10.1016/B978-0-12-821841-9.00007-4 © 2021 Elsevier Inc. All rights reserved.

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The current focus on biohythane is on the reduction of greenhouse gases, effective combustion properties, and optimizing fuel efficiency. The biohythane fuel station will be installed as fossil fuels and used as the main fuel in transport vehicles, with more efficiency and better performance, and with less end cost. However the current independent generation of H2 and CH4 from fossil fuel based material is energy intensive and unsustainable. The generation of biohydrogen and biomethane process can be integrated in one system to form the optimal composition of biohythane by using renewable biomass and wastewater as the initial substrate for this. In the last decade, the bioelectrochemical system (BES) has emerged as a sustainable approach for wastewater treatment (Sevda et al., 2013a, 2016, 2020; Sevda and Abu-Reesh, 2017, 2018; Naha et al., 2020). The biohythane production in BES is a new approach, and it will also provide a flexible control over the H2/CH4 ration in the biohythane generation (Li et al., 2020).

8.2 Main aspects of the biohythane generation in bioelectrochemical system Anaerobic bioprocesses such as BES, fermentative biohydrogen generation, and anaerobic digestion for biomethane production are used for biological wastewater treatment and simultaneously generation of bioelectricity, H2 and CH4, respectively (Premier et al., 2013; Sevda et al., 2014). The BES research has shown that this technology is also suitable for low-strength wastewater (chemical oxygen demand less than 500 mg/L), as compared to anaerobic digestion (Sevda et al., 2018). The anodic chamber of BES is operated under anaerobic condition while the cathodic chamber is operated with oxygen or other terminal electron acceptor (Sevda and Sreekrishnan, 2014). In the microbial fuel cell (MFC) system, wastewater is treated in the anodic chamber and CH4 is also produced in the anodic chamber itself (Durruty et al., 2013). The biohydrogen is produced in the cathodic chamber of the microbial electrolysis cell (MEC) by using CO2 as a substrate, and the wastewater is treated in the anodic chamber of the MEC (Bundhoo, 2017). In the MEC, the electrons generated in the anodic chamber helped to supply less overpotential from an external source, and hence it is more economical. Zhen et al. (2016) investigated the microbial electromethanogenesis concept for methane formation from CO2 on an electrochemically active biofilm on the cathode electrode surface in a MEC reactor. Fig. 8.1 shows a schematic of an MEC system; in this, hybrid-graphite felt was used as cathode electrode. When using a hybrid-graphite felt cathode, methane production of 80.90 mL/L was reported at the potential of 21.4 V with coulombic efficiency of 194.4% (Zhen et al., 2016). The scanning electron microscopy confirms the electroactive cell on the electrode surface (Fig. 8.1). Zhen et al. (2016) also concluded that hybrid-graphite felt acted as artificial pili and helped in direct electron transport between the electrode and cell; cell to cell without any physical contact (Fig. 8.1) resulted in the

8.3 Substrate for biohythane generation

FIGURE 8.1 The schematic of two chambers of MEC used for microbial electromethanogenesis. The electrochemically active microbes on the biocathode perform the electron transfer and convert CO2 to CH4. MEC, Microbial electrolysis cell; GF, graphite felt; PEM, proton exchange membrane. Reproduced with permission from Zhen G., Lu X., Kobayashi T., Kumar G., Xu K., 2016. Promoted electromethanosynthesis in a two-chamber microbial electrolysis cells (MECs) containing a hybrid biocathode covered with graphite felt (GF). Chem. Eng. J. 284, 1146 1155.

performance of the electromethanogenesis process, which was also improved for methane generation from CO2. Luo et al. (2017) developed a new, innovative, large-scale MEC (19 L) and investigated the effect of operational factors (external resistance, anolyte recirculation rate, and hydraulic retention rate [HRT]) on biohythane generation. Fig. 8.2 shows the experimental setup of new innovative MEC for biohythane generation. The optimized process parameters for the present MEC system were external resistance of 1 ohm, anolyte recirculation rate of 0.8 L/min, and HRT of 24 hour (Luo et al., 2017). The obtained biohythane generation of 0.64 6 0.06 L/d with 16.5% H2 proportion was produced in an innovative MEC system. The positive net energy recovered with optimal process parameter was 1.52 6 0.19 kWh/d (Luo et al., 2017). The statistical analysis of variance (ANOVA) indicated that HRT is affecting the influence of organic removal while having less effect on biohythane composition. The ANOVA analysis also indicated that methane generation was influenced by the anolyte recirculation rate while external resistance was affected by the biohydrogen generation in the innovative MEC (Luo et al., 2017).

8.3 Substrate for biohythane generation The ideal substrate for biohythane production is simply an organic compound that is rich in fats, carbohydrates, and proteins. Usually, the commercial production of biohythane

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CHAPTER 8 Potential of high energy compounds: Biohythane

FIGURE 8.2 MEC experimental setup for biohythane production: (A) the components and structure of MEC reactor, (B) the internal structure of cathode electrode and anode electrode (carbon brush), (C) cathode biohythane gas collection setup, and (D) collectively full setup of the whole experiment. MEC, Microbial electrolysis cell. Reproduced with permission from Luo S., Jain A., Aguilera A., He Z., 2017. Effective control of biohythane composition through operational strategies in an innovative microbial electrolysis cell. Appl. Energy [Internet]. 206, 879 886. doi: 10.1016/j.apenergy.2017.08.241.

needs a cheap substrate that kept aside the simple sugars (glucose, maltose, sucrose, and lactose) due to the high cost of the substrates (Lin et al., 2018). The high-cost issues arising with the simple sugars paves the way for researchers to look after polysaccharides (cellulose, hemicellulose, amylose) for a possible substrate for biohythane production. The polysaccharide-bound lignocellulosic biomass seems promising due to high

8.4 Recent progress for biohythane generation in bioelectrochemical

abundance and lower cost but the initial pretreatment and saccharification steps consume more time. Due to the lengthy initial pretreatment step, the overall cost of the process is increased, and resulting from this the polysaccharide-bound lignocellulosic biomass becomes out of the race for the possible future substrate for biohythane production (Kumari and Das, 2019). The search for ideal substrate makes researchers look after different high chemical oxygen demand (COD)-containing wastes such as Municipal Solid Waste (MSW), food, and distillery wastes, which considered a threat to the environment. The results with the CODcontaining wastes were also not successful and resulted in lesser yields even after several pretreatments. The next search for possible substrate eyes on high organic content, volatile suspended solids (VSS)-rich food wastes, high CODand BOD- containing distillery wastes, and palm oil mill effluents (Si et al., 2016a). The research attempted to use lignocellulosic biomass, algal biomass, and agricultural residues as a substrate for biohythane production. Some of the vital research findings toward a possible substrate for biohythane production have been summarized in Table 8.1.

8.4 Recent progress for biohythane generation in bioelectrochemical system In the biohythane generation plant, the most important challenge is to maintain the ratio of H2/CH4 in all types of reactors. The BES reactor provides the better control in the maintain the H2/CH4 ratio in the biohythane production. Li et al. (2020) investigated the effect of an external resistance for maintaining the final ration of H2/CH4 in a dual-chamber BES. Fig. 8.3 shows the innovative threechamber BES reactor with having one anode chamber and different cathode chamber for generation of CH4 and H2 respectively. The anodic chamber reactor volume was double compared to each cathodic chamber of the BES system (Li et al., 2020). Luo et al. (2017) have shown that effective control in biohythane composition can be controlled by the current generation in dual-chamber MEC. In the BES technology more biohydrogen is produced in the MEC mode, while more biomethane is produced in the MFC mode (Luo et al., 2017; Lu et al., 2012; Sevda et al., 2013b; He et al., 2005). Li et al. (2020) demonstrated that when external resistance values were increased from 10 to 330 ohm, the biomethane generation was increased from 25 6 3 to 90 6 5 L/m3/d. In similar conditions the biohydrogen generation was decreased from 173 6 11 to 8 6 2 L/m3/d. The overall H2/CH4 ration in the dual-chamber BES reactor was decreased to 0.09 from 7.14 (Li et al., 2020). Due to changes in the external resistance value in the dual-chamber BES, electron distribution was changed in the cathodic chamber, and it affects for both biohydrogen and biomethane final production rates in the reactor. Depending on the applications of biohythane, the H2/CH4 ration can be obtained in the BES reactor compared to other reactors (Li et al., 2020; Luo et al., 2017).

169

Table 8.1 Use of various reactor types for generation of biohythane. S. No. 1.

2.

3.

4

5

Reactor type

Electrode material, Ion exchange membrane

Substrate composition

Biohythane composition

References

Hydrogen production: 66.7 L/kg of total volatile solid; biogas production: 0.72 m3per per kg of total volatile solid Average specific biohythane production of 0.65 m3/kg of volatile solids

Cavinato et al. (2012)

Biohydrogen yield of 16 mL H2/g VS, biomethane yield of 240 mL CH4/gVS (The biomethane and bio hydrogen content in biogas was 65.21% and 25.33% respectively) Biomethane (2.53 UASB) and 2.54 L/L/d (PBR0); Biohydrogen (2.34 L/L/d in UASB and 1.01 L/L/d in PBR) 2.01 CH4 yield mol/mol; H2/ H2 1 CH4 (v/v): 0.40

Suksong et al. (2015)

Two continuous stirred tank reactors: thermophilic dark fermentation and anaerobic digestion Two continuous stirred tank reactors: a twostage thermophilic and anaerobic digestion process Two-stage solid state anaerobic digestion process

2

Food waste (total volatile solid: 219 g/kg; COD: 257 g/kg ww)

2

Organic fraction of municipal solid waste (total solids: 270 g/kg; COD: 9346 mg/L, 4.2 kg/m3/d)

2

Palm oil mill effluent (total solid: 5.88% (w/w); COD 44.72 g/L; cellulose 4.2%, carbohydrates 50.00 mg/L)

Two-stage fermentation: upflow anaerobic sludge blanket (UASB) and packed bed reactor (PBR) One stage system using UASB and PBR

2

Hydrothermal liquefied cornstalk biomass (COD (mg/L): 76.22; total organic carbon (mg/L): 28.60)

2

Substrate composition (mg/L): K2HPO4 250, KH2PO4 250, MgCl2 300, CoCl2 25, ZnCl2 11.5, CuCl2 10.5, CaCl2 5, MnCl2 15, NiSO4 16, FeCl3 25; glucose NH4Cl was changed to obtain desired OLRs

Giuliano et al. (2014)

Si et al. (2016b)

Si et al. (2016a)

6

An innovative MEC (53.5 3 7.6 3 50.8 cm) system (19 L)

The cathode chamber(s) inside the anode chamber. cathode consisted of five tubular cathode segments, which were constructed with anion exchange membrane (AEM), Carbon brush inside AEM acted as anode electrode

7

An electrochemical HCell type, 360 mL each chamber (anodic and cathodic) A flat plate BES reactor (one anodic chamber and two cathodic chamber)

Cation exchange membrane (9.6 cm2), Carbon cloth anode and cathode electrode (22.5 cm2) Cation exchange membrane (25 cm 3 25 cm), Anode electrode (non-wet proof carbon brush electrode), Cathode electrode (wet proof carbon cloth)

8

Low-strength wastewater contained (per L of DI water): 0.17 g sugar (to achieve a COD concentration of 200 mg/L, granular sucrose); NH4Cl, 0.15 g; NaCl, 0.5 g; MgSO4, 0.015 g; CaCl2, 0.02 g; NaHCO3, 0.5 g; and trace element, 1 mL) Synthetic wastewater, biocathode operational mode

Biohythane generation of 0.64 6 0.06 L/d with 16.5% H2 proportion

Luo et al. (2017)

60.80 CH4 production rate (mmol/m2/d) at cathode potential (vs SHE) of 20.5 V

Li et al. (2019)

Synthetic wastewater (Per L of distilled water): Sodium acetate 1 g/L; NaCl 0.5 g/L; NH4Cl 0.15 g/L; K2HPO4 5.3 g/L; KH2PO4 2.6 g/L; CaCl2 0.02 g/L; NaHCO3 0.1 g/L; MgSO4 0.015 g/L; Trace elements 1 g/L)

CH4 generation rate 25 6 3 to 90 6 5 L/m3/d and H2 production rate 173 6 11 to 8 6 2 L/m3/d were achieved when external resistance was changed from 10 to 330 ohm

Li et al. (2020)

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CHAPTER 8 Potential of high energy compounds: Biohythane

FIGURE 8.3 Experimental schematic of a dual-chamber BES with a CH4-cathode and H2cathode working with similar anode electrode. BES, Bioelectrochemical system; CEM, cation exchange membrane. Reproduced with permission from Li X., Liu G., He Z., 2020. Flexible control of biohythane composition and production by dual cathodes in a bioelectrochemical system. Bioresour. Technol. 295.

8.5 Use of biohythane The usage of hythane from fossil sources for transportation fuel dates back to the 1980s. While going through the properties of methane and hydrogen (Table 8.2), the clean fuel status of methane was limited due to its narrow flammable range and high ignition temperatures, which is responsible for poor combustion efficiency and requirement of high-intensive energy in compressed natural gas (CNG)-powered vehicles (Liu et al., 2013, 2018). The usual composition of biohythane is 5% 10% H2, 60% CH4, and 30% CO2, which necessitates the drawbacks associated with individual fuel properties of biohydrogen and biomethane. The biohythane systems facilitate the combination of hydrogen with methane, which complements the poor fuel properties associated with intact CNG (methane). The complementation helps in enhancing the H/C ratio of intact methane results in lesser GHG emissions. The fuel efficiency of methane, in terms of flammability, also improved by combining hydrogen to methane (through the biohythane system). The combination also facilitates the enhanced ignition properties and easy engine, starting with less energy input than the intact utilization of methane (CNG) as a transportation fuel (Gattrell et al., 2018).

References

Table 8.2 Properties of biomethane and biohydrogen. S. No.

Properties parameter

Biomethane

Biohydrogen

1 2 3 4 5 6 7 8 9 10 11 12 13

Flammability limit in air (% vol.) Molecular weight Ignition temperature (K) Density at standard condition (kg/m3) Burning speed (cm/s) Density at standard condition (kg/m3) Minimum ignition energy (MJ) Mass lower heating value (MJ/kg) Octane number Quenching distance (cm) Adiabatic exponent Stoichiometric air-to-fuel ratio Laminar burning velocity in NTP air (cm/s)

5 15 16 918.00 0.71 37.00 0.71 0.20 50.00 107.50 0.20 1.32 17.19 37 45

4 75 2 858.0 0.09 270.00 0.09 0.02 120.00 . 130 0.06 1.14 34.20 265 325

Modified from Liu Z., Zhang C., Lu Y., Wu X., Wang L., Wang L., et al., 2013. States and challenges for high-value biohythane production from waste biomass by dark fermentation technology. Bioresour. Technol. [Internet] 135, 292 303. http://dx.doi.org/10.1016/j.biortech.2012.10.027.

8.6 Future prospects and concluding remarks The viability of biohythane technology of two-step anaerobic fermentation (TSAF) needs further up-gradation in terms of lower-cost substrate availability, stability concerns of biohydrogen reactors, and amalgamation with BESs toward the creation of biohythane-based biorefinery toward new metabolic products. The amalgamation of biohythane technology through BES and probable effluent treatment using microfluidics platforms solve the scale-up issues associated with the biohythane technology. Biohythane technology can further be upgraded through the creation of potent substrate degradation strains and the utilization of possible mixed microbial consortium associations.

Acknowledgment The authors would like to express gratitude to Research Seed Money Grant provided from National Institute of Technology Warangal, Warangal, India for this work.

References Bolzonella, D., Battista, F., Cavinato, C., Gottardo, M., Micolucci, F., Lyberatos, G., et al., 2018. Recent developments in biohythane production from household food wastes: a review. Bioresour. Technol. [Internet] 257, 311 319. Available from: https://doi.org/ 10.1016/j.biortech.2018.02.092.

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Bundhoo, Z.M.A., 2017. Coupling dark fermentation with biochemical or bioelectrochemical systems for enhanced bio-energy production: a review. Int. J. Hydrogen Energy [Internet] 42 (43), 26667 26686. Available from: https://doi.org/10.1016/j.ijhydene.2017.09.050. Cavinato, C., Giuliano, A., Bolzonella, D., Pavan, P., Cecchi, F., 2012. Bio-hythane production from food waste by dark fermentation coupled with anaerobic digestion process: a long-term pilot scale experience. Int. J. Hydrogen Energy [Internet] 37 (15), 11549 11555. Available from: https://doi.org/10.1016/j.ijhydene.2012.03.065. Durruty, I., Bonanni, P.S., Gonza´lez, J.F., Busalmen, J.P., 2012. Evaluation of potatoprocessing wastewater treatment in a microbial fuel cell. Bioresour. Technol. [Internet] 105, 81 87. [cited May 14, 2013]. Available from: http://www.ncbi.nlm.nih.gov/ pubmed/22178494. Gattrell, M., Gupta, N., Co, A., Cavinato, C., Giuliano, A., Bolzonella, D., et al., 2018. Investigation of microbial biofilm structure by laser scanning microscopy. Bioresour. Technol. [Internet] 5 (4), 9391 9410. Available from: https://doi.org/ 10.1016/j.biortech.2020.122746. Giuliano, A., Zanetti, L., Micolucci, F., Cavinato, C., 2014. Thermophilic two-phase anaerobic digestion of source-sorted organic fraction of municipal solid waste for biohythane production: effect of recirculation sludge on process stability and microbiology over a long-term pilot-scale experience. Water Sci. Technol. 69 (11), 2200 2209. He, Z., Minteer, S.D., Angenent, L.T., 2005. Electricity generation from artificial wastewater using an upflow microbial fuel cell. Environ. Sci. Technol. [Internet] 39 (14), 5262 5267. [cited November 4, 2014]. Available from: http://www.ncbi.nlm.nih.gov/ pubmed/16082955. Kumari, S., Das, D., 2019. Biohythane production from sugarcane bagasse and water hyacinth: a way towards promising green energy production. J. Clean. Prod. [Internet] 207, 689 701. Available from: https://doi.org/10.1016/j.jclepro.2018.10.050. Li, Z., Fu, Q., Kobayashi, H., Xiao, S., Li, J., Zhang, L., et al., 2019. Polarity reversal facilitates the development of biocathodes in microbial electrosynthesis systems for biogas production. Int. J. Hydrogen Energy 26226 26236. Li, X., Liu, G., He, Z., 2020. Flexible control of biohythane composition and production by dual cathodes in a bioelectrochemical system. Bioresour. Technol. 122270. Lin, C.Y., Nguyen, T.M.L., Chu, C.Y., Leu, H.J., Lay, C.H., 2018. Fermentative biohydrogen production and its byproducts: a mini review of current technology developments. Renew. Sustain. Energy Rev. [Internet] 82, 4215 4220. Available from: https://doi. org/10.1016/j.rser.2017.11.001. Liu, Z., Zhang, C., Lu, Y., Wu, X., Wang, L., Wang, L., et al., 2013. States and challenges for high-value biohythane production from waste biomass by dark fermentation technology. Bioresour. Technol. [Internet] 135, 292 303. Available from: https://doi.org/ 10.1016/j.biortech.2012.10.027. Liu, Z., Si, B., Li, J., He, J., Zhang, C., Lu, Y., et al., 2018. Bioprocess engineering for biohythane production from low-grade waste biomass: technical challenges towards scale up. Curr. Opin. Biotechnol. [Internet] 50, 25 31. Available from: https://doi.org/ 10.1016/j.copbio.2017.08.014. Luo, S., Jain, A., Aguilera, A., He, Z., 2017. Effective control of biohythane composition through operational strategies in an innovative microbial electrolysis cell. Appl. Energy [Internet] 206, 879 886. Available from: https://doi.org/10.1016/ j.apenergy.2017.08.241.

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Si, B.C., Li, J.M., Zhu, Z.B., Zhang, Y.H., Lu, J.W., Shen, R.X., et al., 2016b. Continuous production of biohythane from hydrothermal liquefied cornstalk biomass via two-stage high-rate anaerobic reactors. Biotechnol. Biofuels. 9 (1), 1 15. Suksong, W., Kongjan, P., O-Thong, S., 2015. Biohythane production from co-digestion of palm oil mill effluent with solid residues by two-stage solid state anaerobic digestion process [Internet]. Energy Procedia 943 949. Available from: https://doi.org/10.1016/ j.egypro.2015.11.591. Zhen, G., Lu, X., Kobayashi, T., Kumar, G., Xu, K., 2016. Promoted electromethanosynthesis in a two-chamber microbial electrolysis cells (MECs) containing a hybrid biocathode covered with graphite felt (GF). Chem. Eng. J. 284, 1146 1155.

CHAPTER

Biological and chemical remediation of treated wood residues

9

Lais Gonc¸alves da Costa1, Yonny Martinez Lopez1, Victor Fassina Brocco2 and Juarez Benigno Paes1 1

Department of Forest and Wood Science, Federal University of Esp´ırito Santo, Jeroˆnimo Monteiro, Brazil 2 Center for Higher Studies of Itacoatiara, Amazonas State University (CESIT/UEA), Itacoatiara, Brazil

9.1 Introduction Many wood species have a low natural durability, which favors the attack of wood-destroying organisms such as insects, fungi, and marine borers. Thus, preservative treatments with chemicals are necessary to increase the useful life of wood in service. Among the various wood preservatives, those containing heavy metal such as arsenic, chromium, and copper were the most and are still widely used worldwide (Sierra-Alvarez, 2009; Vidal et al., 2015). However, due to the potential risks to public health and the environment, several countries have created restrictions on products that use arsenic in their formulation (Sierra-Alvarez, 2009; Wang et al., 2016). Concerns about the long-term risk of the potential for leaching and environmental contamination by wood treated with several heavy metals are growing. Thus, care must be taken during and after the end of the useful life of the treated wood, as its reuse and inadequate disposal become an aggravating factor. Considering all these concerns, the decontamination of treated wood residues has been widely studied for waste management, whether by chemical extraction methods with organic or inorganic acids, chelating agents, or bioremediation using fungi and bacteria (Sierra-Alvarez, 2009; Wang et al., 2016; Kartal et al., 2014, 2015). Considerable attention has been focused on the biological and chemical remediation of treated wood in recent years, due to public and scientific awareness of the release of heavy metals, such as arsenic, chromium, and copper, which are present in treated wood in its disposal modes, such as disposal in landfills or through burning. Thus, this review brings an approach on the main biological organisms and chemicals studied for the remediation of treated wood residues and their main mechanisms of action. Delivering Low-Carbon Biofuels with Bioproduct Recovery. DOI: https://doi.org/10.1016/B978-0-12-821841-9.00008-6 © 2021 Elsevier Inc. All rights reserved.

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9.2 Environmental risks of treated wood Among the most common wood preservatives (Table 9.1) are those that include inorganic components containing arsenic, boron, chromium, copper, or zinc in their composition, such as acid copper chromate (ACC), ammoniacal copper zinc arsenate (ACZA), chromated copper arsenate (CCA), chromated copper borate (CCB), copper azole (CBA, CA-B), copper citrate (CC), alkaline copper quaternary (ACQ) and zinc borate (ZB) (Vidal et al., 2015; Kartal et al., 2015; Lebow, 2010; American Wood Protection Association AWPA, 2016). Despite the benefits of wood preservatives in the useful life of wood in service, at some point, treated wood reaches the end of its lifetime. Thus, problems start to arise regarding the disposal and reuse of this material, depending on the potential risks related to health and the environment (Wang et al., 2016; Chen, 2015). Most of this waste is improperly disposed, and can cause a series of impacts on the physical environment, which involves the soil, surface, groundwater, and air, as well as on the biological and socioeconomic environment (Preilipper et al., 2016). The disposal of treated wood is even more dangerous due to the presence of heavy metals in the constitution of preservative products. Thus, after the end of the useful life of the treated wood in service, the procedures with residues must involve collection, separation, and remediation for possible reuse or proper disposal (Brazolin et al., 2003). Depending on the specific forms of arsenic, chromium, and copper, these elements can be more or less carcinogenic, mutagenic, and toxic to a wide range of organisms, as well as harmful to the environment (Hingston et al., 2001; Kartal and Imamura, 2003). Environmental concerns are growing mainly due to the increase of the amount of this waste and the volume of Table 9.1 Concentrations of active inorganic ingredients (%) of some commonly used wood preservatives. Wood preservative CCA-C ACZA ACQ-B CA-B ACC ZB CBA-A CC CCB 

Concentrations of active inorganic ingredients (%) As2O5

CrO3/K2Cr2O7

CuO/Cu /CuSO4

34.0 25.0

47.5

68.2

18.5 50.0 66.7 96.0 31.8

38.5

49.0 62.0 35.8

B2O3/H3BO3

25.0

32.8

Preservative containing the second compound in respective column Preservative containing the third compound in respective column



ZnO

48.2

22.4

9.2 Environmental risks of treated wood

treated wood residues that will be removed from service will increase significantly in several countries (Janin et al., 2012; Coudert et al., 2013). Treated wood poles removed from service are an example of this type of waste (Fig. 9.1). Due to its high level of initial preservative retention, more attention should be given to this material. Some authors cite the possibility of reusing these residues, for example, in rural buildings; however, there are concerns about the danger of improper use, storage, and the efficiency of this product after years of weathering (Cooper et al., 1996, 2001; Ferrarini et al., 2015). Localized leaching of arsenic, chromium, and copper from treated wood posts and poles occurs in the surrounding soil, and the arsenic from soils contaminated with CCA appears to have a certain bioavailability compared to other anthropogenic sources (Townsend et al., 2003; Juhasz et al., 2011; Punshon et al., 2017). In Australia, soils in contact with treated vineyard posts exceeded, in some cases, the recommended guidelines for chromium and arsenic in agricultural soils (Vogeler et al., 2005).

FIGURE 9.1 Residues of CCA-treated wood poles removed from service. CCA, Chromated copper arsenate.

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Evidence of soil contamination with heavy metals from treated wood has been reported in the literature. Soils around treated wood structures have been found to contain 8 36 times more arsenic than permitted target levels (Townsend et al., 2003). Also, it was found that treated wood used to construct raised garden beds significantly influence the plant uptake of heavy metals, and some of the vegetable crops may not be safe for sustained consumption (Rahman et al., 2004). In the same sense, places contaminated with industrial residues showed an absorption of arsenic (As) in the soil, roots, and aerial parts of some plants, and some cultures exceeded the safety levels limited by World Health Organization (WHO) and Food and Agriculture Organization of the United Nations (FAO) (Lim and McBride, 2015).

9.3 Remediation and recovery of treated wood 9.3.1 Bioremediation The use of microorganisms as an alternative for the removal of chemical components in the treated wood residues has been widely reported. Fungi belonging to the phylum Ascomycota and Basidiomycota showed potential to remediate a wide variety of organic pollutants, such as Phanerochaete chrysosporium, Nematoloma spp., Pleurotus spp., and Trametes spp., which mineralize compounds such as chloroaromatics, aromatic hydrocarbons, and trinitrotoluene (Harms et al., 2011). The majority of mentioned fungi have specific enzymes, as laccases for example, which can be used in the treatment of waste, such as the discoloration and detoxification of effluents containing dyes from the fabric industries and in the effluents of the bleaching plants of the pulp and paper industries (Harms et al., 2011). Regarding chemically treated wood, research that evaluates the ability of fungi and bacteria to remove heavy metals present in the composition of the CCA shows promising results, mainly due to the high efficiency compared to the processes of chemical remediation. As in chemical extraction, the bioremediation of treated wood may depend on a number of variables, such as the microorganism species, preservative treatment used, wood geometry, culture conditions, and the duration of extraction (Clausen and Lebow, 2011). Some microorganisms have the ability of oxidizing or reducing arsenic, chromium, and copper to water-soluble forms, which can be removed from treated wood (Kartal and Imamura, 2003; Felton and De Groot, 1996). However, as known, some species of fungi can degrade and compromise the major constituents of wood such as cellulose, hemicellulose, and lignin (Chang et al., 2012; Xie et al., 2008). Several studies related the production of organic acids by fungi during fermentation plays a key role in the removal of heavy metals from treated wood, mainly copper, chromium, and arsenic. The brown rot fungi, Fomitopsis palustris, Coniophora puteana, and Laetiporus sulphureus, and the mold fungus Aspergillus niger, has the

9.3 Remediation and recovery of treated wood

ability to produce oxalic acid in different concentrations and remediate CCA-treated wood (Kartal et al., 2004a,b). According to these studies, the exposure of CCAtreated wood chips to the fungus A. niger for 10 days promoted a 97% decrease in the amount of arsenic. As mentioned for A. niger, other mold fungi such as Aureobasidium pullulans, Gliocladium virens, Penicillium funiculosum, Rhizopus javanicus, Ceratocystis pilifera, Alternaria alternata, Trichoderma viride, and Cladosporium herbarum showed some ability to produce oxalic acid and decrease the concentrations of arsenic, chromium, and copper. Regarding this approach, A. niger produced the highest amount of oxalic acid (4.6 g L21) among the fungi tested and had the best performance for chromium removal (51%), while P. funiculosum removed 88% of arsenic and T. viride completely removed copper (Kartal et al., 2006). Oxalic acid has played an important role in the partial solubilization of the insoluble metallic compounds fixed in the CCA-treated wood. Some fungi are able to secrete oxalic acid in various concentrations in the culture broth, such as A. niger, which produced a concentration of 34 kg m23 of that acid, and also citric acid and gluconic acids (Cameselle et al., 1998). For organic acid production using rot fungi (F. palustris, C. puteana, and L. sulphureus) in fermentation broths to remediate treated wood, oxalic acid was the only one detected (Kartal et al., 2004b). Oxalic acid has been directly related to the deterioration process of wood by brown rot fungi, as well as in the reduction of pH and in the catalyzed hydrolysis of the wood substrate. As shown by the same study, F. palustris exhibited the highest levels of oxalic acid accumulation per liter of fermentation broth and the highest pH reduction, from 5.6 to 3.6. F. palustris, C. puteana, and L. sulphureus produced 4.2, 0.3, and 3.2 g L21 oxalic acid, respectively, after a 10-day fermentation (Kartal et al., 2004b). As shown by fungi, bacteria also have the potential to be used in the bioremediation of treated wood. Some bacteria are able to release heavy metals from CCAtreated wood, as well as metabolizing material chemicals such as benzene, dioxin, dichlorodiphenyltrichloroethane (DDT), polychlorinated biphenyl (PCB), styrene, xylene, oil, gasoline, pentachlorophenol, creosote, and other materials such as tires and concrete (Clausen, 2000b). Among 28 bacterial species investigated, three isolates were able to remove 98% of the chromium: Acinetobacter calcoaceticus, Aureobacterium esteroaromaticum, and Klebsiella oxytoca (Clausen, 2000b). Since some bacteria are extremely tolerant of toxic metals, bacterial fermentation becomes a possible method for removing heavy metals from treated wood. The mechanism for bacterial removal of metals occurs by converting CCA elements into their water-soluble form. Once converted, arsenic, chromium, and copper can be removed from the wood by the water washing process (Kartal and Clausen, 2001). Regarding capacity of the bacterium Bacillus licheniformis in removing toxic metals from CCA-treated wood, when oxalic acid remediation was combined with B. licheniformis remediation, 90%, 79%, and 62% of arsenic, chrome, and copper were removed, respectively (Kartal and Clausen, 2001). According to the

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CHAPTER 9 Biological and chemical remediation of treated wood residues

same study, when the remediation was performed only with oxalic acid, 74%, 65%, and 23% of arsenic, chrome, and copper were removed, respectively. The approach of combined biological and chemical extraction is highly recommended to improve the metal removal from treated wood (Kartal and Imamura, 2003; Clausen, 2000a; Helsen and Van Den, 2005). These authors pointed out that the biological extraction of residues from treated wood allows for almost complete removal and that the combination of extraction with solvents is used to extract as much of the toxic metals as possible. The extraction of oxalic acid followed by bioremediation with a metal-tolerant bacterium was found to remove 80 100% of the CCA components in liquid culture medium, which could be recovered and reused for production of particleboard panels (Wang et al., 2016; Kartal and Clausen, 2001). Lactic acid producing bacteria, such as, Lactobacillus bulgaricus, Lactobacillus acidophilus, Lactobacillus plantarum, and Streptococcus thermophilus also have the ability to remove metal components from CCA-treated wood (Chang et al., 2012). The rates of extraction of heavy metals for L. bulgaricus and S. thermophilus were the highest and are related to the highest concentrations of pyruvic and lactic acid produced by them (Chang et al., 2012). In addition, according to the aforementioned authors, the extraction of CCA elements when performed with a mixed culture (L. bulgaricus and S. thermophilus) is more effective. The extraction rate of arsenic, chrome, and copper in the mixed culture was improved from 66% to 73%, from 45% to 55%, from 55% to 65%, respectively. However, some research emphasizes the limitations of the use of biological remediation. According to this research, despite being technically feasible, it is a relatively slow method and has a high cost when compared to disposal in landfills due to the high cost with nutrients from the culture medium (Helsen and Van Den, 2005; Clausen, 2004). Malted barley, an abundant byproduct of beer production, was used as an alternative to the commercial culture medium to verify an improvement in chemical remediation (oxalic acid) preceded by biological remediation (Clausen, 2004). However, the culture media with malted barley were not as effective in removing metal as the commercial ones, with 78%, 54%, and 21% removal of arsenic, chrome, and copper, respectively, in the culture media with malted barley. While in commercial media the removal was 93%, 97%, and 78% for arsenic, chrome, and copper, respectively. Thus, according to same study, the two-stage extraction process with oxalic acid and bacterial culture, although more expensive, is the most effective method of removing metals. Researchers have also been dedicated to assessing the efficiency of other materials of biological origin in removing arsenic, chromium, and copper from CCA-treated wood. Activated carbon, Sugi (Cryptomeria japonica) wood charcoal, commercial black tea leaf extract, pine and oak bark, bakery yeast, apple and orange peelings, fungus biomass (Fomes fomentarius), pine cones, barley waste, and corn cobs were used to determine the efficiency in removing heavy metals from CCA-treated wood (Kartal et al., 2008).

9.3 Remediation and recovery of treated wood

In the aforementioned study, CCA-treated wood extraction was more effective with the use of activated charcoal, C. japonica wood charcoal, and orange peel. In general, chrome was the most resistant to remove the different types of extractors, and copper was the easiest to remove. Thus, using a greater extraction of arsenic, chromium, and copper is advisable to carry out useful research or the double remediation process using chemical or biological remediation (fungi and bacteria). Table 9.2 shows different types of bacteria and fungi used in some studies for the removal of chemical components from CCA in treated wood.

9.3.2 Mechanisms used by fungi in the remediation process It is known that several biological organisms are used to remediate materials contaminated with heavy metals, including species of bacteria and fungi, and most of these have adapted to the contaminated environment, or naturally have specific mechanisms for remediation (Liu et al., 2017). Thus, some fungi are able to support and detoxify metal ions by mechanisms such as active accumulation, intracellular and extracellular precipitation, and change in valence state; therefore, they are potential biocatalysts for the Table 9.2 List of studies on the removal of arsenic, chromium, and copper from wood residues treated with CCA, using bacteria and fungi. Rmemoval of elements (%) Microorganism

Specie

Extraction time

As

Cr

Cu

Source

Bacterium

Bacillus licheniformis

7 days

45

8

93

Clausen (2000b)

Bacterium

Aureobacterium barkeri

7 days

37

68

50

Clausen (2000b)

Bacterium

Acinetobacter calcoaceticus

7 days

48

97

25

Clausen (2000b)

Bacterium

Lactobacillus bulgaricus

4 days

66

42

55

Chang et al. (2012)

Bacterium

Streptococcus thermophilus

4 days

62

45

51

Chang et al. (2012)

Fungus

Coniophora puteana

10 days

18

19

67

Kartal et al. (2004b)

Fungus

Laetiporus sulphureus

10 days

85

69

50

Kartal et al. (2004b)

Fungus

Fomitopsis palustris

10 days

100

87

72

Kartal et al. (2004b)

Fungus

Aspergillus niger

10 hours

97

55

49

Kartal et al. (2004a)

Fungus

A. niger

10 days

78

Kartal et al. (2015)

Fungus

Irpex lacteus

10 days

61

Kartal et al. (2015)

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CHAPTER 9 Biological and chemical remediation of treated wood residues

bioremediation of heavy metals because they are able to absorb them in mycelia and spores (Kartal and Imamura, 2003; Ojuederie and Babalola, 2017). Although heavy metals such as arsenic, chromium, and copper cannot be destroyed from the medium or transformed into nontoxic forms, they can be reduced or oxidized to water-soluble forms. Thus, the use of microorganisms for this purpose has been presented as an environmentally friendly and low-cost technology (Kartal and Imamura, 2003; Liu et al., 2017). The bioremediation of treated wood involves complex biological, chemical, and physical reactions that are capable of immobilizing or transforming heavy metals in less toxic forms (Kartal and Imamura, 2003). Other authors treat bioremediation as a challenging process because the presence of heavy metals can negatively affect microbial activities, such as the production of adenosine triphosphate (ATP), carbon mineralization, displacement of the community, and enzymatic function (Liu et al., 2017). A wide variety of microorganisms promote leaching of metals from CCAtreated wood. This mechanism is linked to the ability of microorganisms to oxidize or reduce arsenic, chromium, and copper in water-soluble forms, which can be removed from treated wood (Kartal and Imamura, 2003). Some classes of fungi play an important role in the remediation of CCA-treated wood, where heavy metals can be transformed by the enzymatic systems of these fungi (Kartal and Imamura, 2003). Fungal tolerance to heavy metals can be promoted by several mechanisms, such as accumulation or biosorption of heavy metals by cell wall components and extracellular materials, or precipitation by secreted metabolites, such as enzymes or acids (Kartal and Imamura, 2003; Liu et al., 2017). For example, the tolerance of some decay fungi to the copper element has been linked to the amount of oxalic acid produced by the fungi. Oxalate is an agent that penetrates the structure of the cell wall of the wood and can work together with metals in the depolymerization reactions of the components present in the wood cells (Kartal and Imamura, 2003). Oxalic acid can be produced in a biotechnological process because some fungi are able to secrete it in various concentrations within the culture medium. According to the aforementioned authors, the removal of arsenic, chromium, and copper from CCA-treated wood increased significantly during extraction with oxalic acid. Brown rot fungi are reported to produce high levels of oxalic acid and other polycarboxylic acids. Some brown rot fungi with prominence in the literature for the production of oxalic acid are F. palustris, C. puteana, and L. sulphureus. Mold fungi like A. niger and Penicillium spp. are also reported as potential for bioremediation according to the oxalic acid production (Kartal and Imamura, 2003; Cameselle et al., 1998). The most efficient producers of oxalic acid and, consequently, the most tolerant fungi also include the genus Antrodia (Humar et al., 2004). The fungus A. niger produces not only oxalic acids but also citric and gluconic acids, according to the conditions of the environment. According to the aforementioned authors, the synthesis of these acids is performed through the Krebs cycle, where gluconic acid is produced by the oxidation of glucose in a rapid reaction catalyzed by

9.3 Remediation and recovery of treated wood

glucose oxidase. The glycolysis pyruvate is transformed into oxalacetate, the latter being hydrolyzed to oxalate for the production of oxalic acid (Cameselle et al., 1998). White rot fungi are also useful for bioremediation. These fungi can transform or mineralize a series of organic chemicals present in the wood, such as creosote and pentachlorophenol (Mileski et al., 1988). Also, the white rot fungus P. chrysosporium was able to mineralize more than 50% of pentachlorophenol. It was concluded that low levels of CCA retention were necessary to prevent the growth of fungi Irpex lacteus and Trametes versicolor that cause white rot, compared to the growth of brown rot fungi (Jusoh and Kamdem, 2001). Although they are generally less tolerant to copper-based wood preservatives than brown rot fungi, some white rot fungi are capable of deteriorating CCA-treated wood. White rot fungi can concentrate the metals removed from the substrate in their mycelium by biosorption. In addition, their fruiting bodies that grow on the wood can accumulate the copper on the treated wood (Kartal and Imamura, 2003). Another mechanism reported for white rot fungi is hemicellulolytic and ligninolytic enzymes secreted that play a role in the removal or deterioration of heavy metals. In some cases, the fungal cell structure has the ability to absorb heavy metals from various media containing heavy metal ions. Biosorption of heavy metals by fungi occurs as a result of ionic interactions and a complex formation between metal ions and functional groups present on the surface of fungi cells. The functional groups responsible for the biosorption of heavy metals are phosphate, carboxyl, amine, and amide (Kartal and Imamura, 2003; Kapoor and Viraraghavan, 1997). At the same time, the bioremediation process does not depend only on the microorganisms used. The conditions for the remediation must be carried out with the objective of microbial growth, from the supply of inorganic nutrients, oxygen, humidity, and adequate temperature. The process of capturing metal, for example, is complex and dependent on the chemistry of metal ions, specific surface properties of organisms, cell physiology, and physical conditions, such as pH, temperature, and metal concentration of the medium (Kartal and Imamura, 2003). Some authors have described information in the literature to improve the biosorption of preservative products, such as by changing pH levels in aqueous solutions of fungi. The biosorption capacity of A. niger in the removal of pentachlorophenol from aqueous solutions has been reported to be pH dependent (Ojuederie and Babalola, 2017). The same authors reported several results that highlight the importance of using the appropriate pH for the best performance of the microorganisms used in bioremediation. Temperature has a crucial effect on the heavy metal bioremediation system. The solubility of heavy metals increases with increasing temperature, which improves the bioavailability of heavy metals; however, this increase must occur within an appropriate range to enhance the metabolism of microorganisms and enzymatic activity, accelerating the bioremediation process (Liu et al., 2017). According to the same authors, the mechanisms of action of fungi for remediation are highly dependent on the characteristics of the environment; therefore, the appropriate conditions must be sought for the process to reach its maximum efficiency.

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The use of microorganisms for the remediation of treated wood show high efficiency compared to chemical extraction processes. On the other hand, more information about the mechanisms of detoxification and the interaction with the characteristics of the process are necessary to evaluate the biodeterioration capacity and the development of new technologies.

9.3.3 Chemical remediation The main concerns with the use of treated wood residues have been the possible release of its constituent elements, such as arsenic, chromium, and copper, during the management of these residues, whether for reuse, recycling, landfill, or burning (Kartal and Imamura, 2003). Several studies have been carried out with chemical extraction to remove arsenic, chromium, and copper from CCA-treated wood (Janin et al., 2009, 2011; Kartal, 2003; Nanseu-Njiki et al., 2007). Chemical extraction with acids and oxides reverses the process of fixing CCA in wood, increasing the solubility of metals in the extraction solution (Clausen and Lebow, 2011). The objective of the remediation process of treated wood involves the extraction of arsenic, chromium, and copper, and, consequently, its transfer from wood residues to a solution, aiming to alleviate these environmental concerns for the management of treated wood and to ensure the safety of workers involved in the management of this waste (Kartal and Imamura, 2003; Clausen, 2000a). Basically, the chemical remediation process takes place as follows: the treated wood residues are reduced to sawdust or chips to obtain smaller particles, which can be easily washed by other chemicals used for extraction. From this point, the treated wood particles undergo a washing process with an extraction solution for a certain time, resulting in free particles of the compounds destined for removal. The result of the removal process is a washing solution (Fig. 9.2) containing the heavy metal compounds removed, which can be concentrated and recovered for other uses (Nanseu-Njiki et al., 2007). Certain factors such as size of the wood particles, diffusion of the chemical in the wood, concentration of the extraction solution, pH, temperature, and extraction time are determinants for the extraction of chemicals from treated wood residues. On the other hand, prolonged exposure to strong acids can damage fiber integrity, and therefore, exposure for the shortest possible time is desirable (Kartal and Imamura, 2003; Clausen, 2000a). In addition, the energy expended in the process and the health risk during the preparation of the chips or sawdust used in the remediation processes must be taken into account. In preparing sawdust, the CCA-treated wood particles released into the air can be highly dangerous to the health of workers. The use of chips in the remediation process is more realistic compared to sawdust, since less energy is spent for its preparation with less risk to health (Velizarova et al., 2002). Most of the studies cited suggest the viability of the various organic acids for removing arsenic, chromium, and copper from treated wood, such as citric, acetic, formic, oxalic, fumaric, tartaric, gluconic, malic, sulfuric, hydrochloric, nitric,

9.3 Remediation and recovery of treated wood

FIGURE 9.2 Washing solution in different acids used for chemical remediation of treated wood. EDTA, Ethylenediamine tetraacetic acid.

phosphoric acid, and ethylenediamine tetraacetic acid (EDTA). Other authors extracted about 95% of the chemical components of the treated wood using hydrogen peroxide in a short time of extraction (6 hours) (Kazi and Cooper, 2006). According to the aforementioned authors, acid extraction is efficient in removing the components of preservative products and factors such as extraction time, temperature adjustment, concentration of the extraction solution, and other variables must be tested to improve the process. Another possibility mentioned to increase the extraction efficiency would be the use of a double extraction for the same sample of the treated wood, with the combination of several acids or the combination with biological extraction methods (Kartal and Imamura, 2003). On the other hand, the high amount of chemicals used for the extraction and the various stages to guarantee the removal of the active ingredients are pointed out as a disadvantage of the process. In addition, the technology for recovering the extracted products is not yet fully understood (Helsen and Van Den, 2005). The costs of chemicals used in remediation must be considered, as most reagents are expensive and can affect the process development on a large scale. Sulfuric and oxalic acids showed good extraction yields and low costs in the order of 9.1 US$ ttw21 and 84 US$ ttw21 (ttw: ton of treated wood), respectively, when compared with other chemical such as phosphoric acid, hydrogen peroxide, and EDTA (Janin et al., 2009). Also, EDTA and phosphoric acid are not efficient in solubilizing the chromium of treated wood; together with this, they have higher costs.

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Another problem with chemical remediation procedures is the guarantee of equivalent extraction efficiency for all metallic components of CCA or other preservatives. It is noted, for example, that some extraction solutions are efficient to eliminate As and Cr, while others are more selective for Cu (Clausen and Lebow, 2011). For the combination of acetic, oxalic, and phosphoric acids in the extraction of the constituent metals of CCA-treated Pinus sp., the highest recovery rate was found for the combination of phosphoric and acetic acid in the concentrations of 2%, 75%, and 0.50%, respectively, at 130 C. The extraction values were approximately 100.0%, 96.7%, and 98.6% for As, Cr, and Cu, respectively (Hse et al., 2013). Chelating agents also have the ability to separate metals from contaminated waste by forming soluble metal-chelate complexes. One of the most efficient chelating agent for the extraction of CCA in treated wood was [S, S]-ethylenediaminedisuccinic acid (EDDS), with maximum extraction efficiency (pH 4) of 58%, 63%, and 93% for As, Cr, and Cu, respectively (Chang et al., 2013). Wood vinegar was also tested as an alternative chemical to evaluating the capacity to extract arsenic, chromium, and copper from CCA-treated wood, which obtained satisfactory results. The higher concentration of wood vinegar and the longer time and temperature extraction conditions increased the extraction efficiency, resulting in the removal of up to 92.7%, 86.3%, and 95.7% of As, Cr, and Cu in sawdust from CCA-treated wood, respectively (Choi et al., 2012). According to the authors, the extraction of the elements present in the CCA was significantly affected by the concentration of wood vinegar and the extraction conditions, such as time and temperature. Table 9.3 shows some products used in

Table 9.3 Removal of arsenic, chromium, and copper from CCA-treated wood residues, using organic acids and other substances. Removal of elements (%) Solution

Extraction time (h)

As

Cr

Cu

Source

Bioxalate Bioxalate Bioxalate Sulfuric acid Oxalic acid Oxalic acid Oxalic acid Sulfuric acid Sulfuric acid Sulfuric acid Sulfuric acid EDTA EDTA

6 6 24 22 6 22 24 1 6 6 6 6 24

89

88

67

48

80

61

97 96 87 93

88 78 70 87

94 94 95 100 47 49 47 96 91 76 98 92 97

Kakitani et al. (2006) Kartal et al. (2014) Kartal et al. (2014) Janin et al. (2009) Kartal et al. (2014) Janin et al. (2009) Kartal et al. (2014) Coudert et al. (2013) Coudert et al. (2014) Coudert et al. (2014) Euflosino (2015) Kartal et al. (2014) Kartal et al. (2014)

References

several studies to remove the chemical components of the CCA and other products in the treated wood.

9.4 Concluding remarks Overall, this chapter reviews the possible environmental impacts, microorganisms, and chemicals for treated wood remediation, summarizing the mechanisms and factors that affect the efficiency of remediation of treated wood waste. The total valorization of wood residues combined with practices to minimize the environmental impact of hazardous waste tends to become increasingly urgent. Thus, it is necessary to conduct future research to improve economically viable methods of remediation and for commercial scale use, since most of the studies mentioned reveal only laboratory methods. In addition, future studies should verify the properties of the recovered particles to test the feasibility of its reuse for the manufacture of other wood byproducts, as well as the purpose of energy use.

References American Wood Protection Association (AWPA), 2016. AWPA Book of Standards. AWPA, Birmingham, AL. Brazolin, S., Romagnano, L.F., Silva, G.A., 2003. Madeira preservada no ambiente constru´ıdo: cena´rio atual e tendeˆncias. In: III ENECS - Encontro Nacional sobre Edif. e comunidades sustenta´veis, Sa˜o Carlos, 15 p. Cameselle, C., Bohlmann, J.T., Nu´n˜ez, M.J., Lema, J.M., 1998. Oxalic acid production by Aspergillus niger. Bioprocess. Eng. 19, 247 252. Available from: https://doi.org/ 10.1007/PL00009017. Chang, Y., Choi, D., Kikuchi, S., 2012. Enhanced extraction of heavy metals in the twostep process with the mixed culture of Lactobacillus bulgaricus and Streptococcus thermophilus. Bioresour. Technol. 103, 477 480. Available from: https://doi.org/10.1016/j. biortech.2011.09.059. Chang, F.C., Wang, Y.N., Chen, P.J., Ko, C.H., 2013. Factors affecting chelating extraction of Cr, Cu, and As from CCA-treated wood. J. Environ. Manag. 122, 42 46. Available from: https://doi.org/10.1016/j.jenvman.2013.02.012. Chen, S.A., 2015. Green and novel technology for recovering copper and wood from treated wood waste Part 1. In: Annual Meeting of International Research Group on Wood Protection, 46, Vin˜a del Mar. Choi, Y.S., Ahn, B.J., Kim, G.H., 2012. Extraction of chromium, copper, and arsenic from CCA-treated wood by using wood vinegar. Bioresour. Technol. 120, 328 331. Available from: https://doi.org/10.1016/j.biortech.2012.06.045. Clausen, C.A., 2000a. CCA removal from treated wood using a dual remediation process. Waste Manag. Res. 18, 485 488. Available from: https://doi.org/10.1177/0734242X0001800510. Clausen, C.A., 2000b. Isolating metal-tolerant bacteria capable of removing copper, chromium, and arsenic from treated wood. Waste Manag. Res. 18, 264 268.

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Clausen, C.A., 2004. Improving the two-step remediation process for CCA-treated wood: Part II. Evaluating bacterial nutrient sources. Waste Manag. 24, 407 411. Available from: https://doi.org/10.1016/j.wasman.2003.11.007. Clausen, C.A., Lebow, S.T., 2011. Reuse and disposal. In: Morrell, C.M., Brooks, J.J., Davis, K.M. (Eds.), Managing Treated Wood in Aquatic Environments. Forest Products Society, Madison, WI, pp. 435 449. Cooper, P., Ung, T., Aucoin, J.-P., Timusk, C., 1996. The potential for re-use of preservativetreated utility poles removed from service. Waste Manag. Res. 14, 263 279. Cooper, P.A., Jeremic, D., Taylor, J.L., 2001. Residual CCA levels in CCA-treated poles removed from service. For. Products J. 51, 58 62. Coudert, L., Blais, J.F., Mercier, G., Cooper, P., Gastonguay, L., Morris, P., et al., 2013. Pilot-scale investigation of the robustness and efficiency of a copper-based treated wood wastes recycling process. J. Hazard. Mater. 261, 277 285. Available from: https://doi.org/10.1016/j.jhazmat.2013.07.035. Coudert, L., Blais, J.F., Mercier, G., Cooper, P., Janin, A., Gastonguay, L., 2014. Demonstration of the efficiency and robustness of an acid leaching process to remove metals from various CCA-treated wood samples. J. Environ. Manag. 132, 197 206. Available from: https://doi.org/10.1016/j.jenvman.2013.11.028. Euflosino, A.E.R., 2015. Ana´lise qu´ımica da madeira tratada com CCA - C, do carva˜o vegetal e do licor pirolenhoso (M.Sc. thesis). Universidade Federal do Esp´ırito Santo, Jeroˆnimo Monteiro, ES. Felton, C.C., De Groot, R.C., 1996. The recycling potential of preservative-treated wood. For. Products J. 46, 37 46. Ferrarini, S.F., Santos, H.S., Miranda, L.G., Azevedo, C.M.N., Maia, S.M., Chaves, E.S., et al., 2015. Determination of As, Cr, and Cu concentrations in CCA-treated wood poles using acid decomposition in closed flasks, oven heating, and ICP-MS analysis. At. Spectrosc. 36, 187 195. Harms, H., Schlosser, D., Wick, L.Y., 2011. Untapped potential: exploiting fungi in bioremediation of hazardous chemicals. Nat. Rev. Microbiol. 9, 177 192. Available from: https://doi.org/10.1038/nrmicro2519. Helsen, L., Van Den, E., 2005. Bulck, Review of disposal technologies for chromated copper arsenate (CCA) treated wood waste, with detailed analyses of thermochemical conversion processes. Environ. Pollut. 134, 301 314. Available from: https://doi.org/ 10.1016/j.envpol.2004.07.025. Hingston, J.A., Collins, C.D., Murphy, R.J., Lester, J.N., 2001. Leaching of chromated copper arsenate wood preservatives: a review. Environ. Pollut. 111, 53 66. Available from: https://doi.org/10.1016/S0269-7491(00)00030-0. Hse, C.Y., Shupe, T.F., Yu, B., 2013. Rapid microwave-assisted acid extraction of southern pine waste wood to remove metals from chromated copper arsenate (CCA) treatment. Holzforschung 67. Available from: https://doi.org/10.1515/hf-2012-0073. ˇ Humar, M., Bokan, M., Amartey, S.A., Sentjurc, M., Kalan, P., Pohleven, F., 2004. Fungal bioremediation of copper, chromium and boron treated wood as studied by electron paramagnetic resonance. Int. Biodeterior. Biodegrad. 53, 25 32. Available from: https://doi.org/10.1016/J.IBIOD.2003.08.001. Janin, A., Blais, J.-F., Mercier, G., Drogui, P., 2009. Optimization of a chemical leaching process for decontamination of CCA-treated wood. J. Hazard. Mater. 169, 136 145. Available from: https://doi.org/10.1016/j.jhazmat.2009.03.064.

References

Janin, A., Coudert, L., Riche, P., Mercier, G., Cooper, P., Blais, J.-F., 2011. Application of a CCA-treated wood waste decontamination process to other copper-based preservative-treated wood after disposal. J. Hazard. Mater. 186, 1880 1887. Available from: https://doi.org/10.1016/J.JHAZMAT.2010.12.094. Janin, A., Coudert, L., Blais, J.F., Mercier, G., Cooper, P., Gastonguay, L., et al., 2012. Design and performance of a pilot-scale equipment for CCA-treated wood remediation. Sep. Purif. Technol. 85, 90 95. Available from: https://doi.org/10.1016/j. seppur.2011.09.052. Juhasz, A.L., Weber, J., Smith, E., 2011. Predicting arsenic relative bioavailability in contaminated soils using meta analysis and relative bioavailability-bioaccessibility regression models. Environ. Sci. Technol. 45, 10676 10683. Available from: https://doi.org/ 10.1021/es2018384. Jusoh, I., Kamdem, D.P., 2001. Laboratory evaluation of natural decay resistance and efficacy of CCA-treated rubberwood (Hevea brasiliensis Muell. Arg.). Holzforschung 55. Available from: https://doi.org/10.1515/HF.2001.041. Kakitani, T., Hata, T., Kajimoto, T., Imamura, Y., 2006. A novel extractant for removal of hazardous metals from preservative-treated wood waste. J. Environ. Qual. 35, 912. Available from: https://doi.org/10.2134/jeq2005.0318. Kapoor, A., Viraraghavan, T., 1997. Heavy metal biosorption sites in Aspergillus niger. Bioresour. Technol. 61, 221 227. Available from: https://doi.org/10.1016/S0960-8524 (97)00055-2. Kartal, S.N., 2003. Removal of copper, chromium, and arsenic from CCA-C treated wood by EDTA extraction. Waste Manag. 23, 537 546. Available from: https://doi.org/ 10.1016/S0956-053X(02)00143-5. Kartal, S.N., Clausen, C.A., 2001. Leachability and decay resistance of particleboard made from acid extracted and bioremediated CCA-treated wood. Int. Biodeterior. Biodegrad. 47, 183 191. Available from: https://doi.org/10.1016/S0964-8305(01)00047-6. Kartal, S.N., Imamura, Y., 2003. Chemical and biological remediation of CCA-treated waste wood. Wood Res. 90, 111 115. Available from: https://doi.org/10.1002/ chin.200425291. Kartal, S.N., Kakitani, T., Imamura, Y., 2004a. Bioremediation of CCA-C treated wood by Aspergillus niger fermentation. Holz Roh Werkst. 62, 64 68. Available from: https:// doi.org/10.1007/s00107-003-0445-y. Kartal, S.N., Munir, E., Kakitani, T., Imamura, Y., 2004b. Bioremediation of CCA-treated wood by brown-rot fungi Fomitopsis palustris, Coniophora puteana, and Laetiporus sulphureus. J. Wood Sci. 50, 182 188. Available from: https://doi.org/10.1007/s10086003-0544-8. Kartal, S.N., Katsumata, N., Imamura, Y., 2006. Removal of copper, chromium, and arsenic from CCA-treated wood by organic acids released by mold and staining fungi. For. Products J. 56, 33 37. Kartal, S.N., Kose, C., Tarakanadha, B., Imamura, Y., 2008. Adsorption of copper, chromium, and arsenic from chromated copper arsenate (CCA) treated wood onto various adsorbents. Open. Waste Manag. J. 1, 11 17. Available from: https://doi.org/10.2174/ 1876400200801010011. Kartal, S.N., Terzi, E., Woodward, B., Clausen, C.A., Lebow, S.T., 2014. Chemical remediation of wood treated with micronised, nano or soluble copper preservatives. Holzforschung 68, 831 837. Available from: https://doi.org/10.1515/hf-2013-0212.

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Kartal, S.N., Terzi, E., Yılmaz, H., Goodell, B., 2015. Bioremediation and decay of wood treated with ACQ, micronized ACQ, nano-CuO and CCA wood preservatives. Int. Biodeterior. Biodegrad. 99, 95 101. Available from: https://doi.org/10.1016/j.ibiod.2015.01.004. Kazi, F.K.M., Cooper, P.A., 2006. Method to recover and reuse chromated copper arsenate wood preservative from spent treated wood. Waste Manag. 26, 182 188. Available from: https://doi.org/10.1016/j.wasman.2004.12.025. Lebow, S.T., 2010. Wood preservation. Wood Handbook - Wood as an Engineering Material. General. Technical Report FPL-GTR-190. U.S. Department of Agriculture, Forest Service, Forest Products Laboratory, Madison, WI. Lim, M.P., McBride, M.B., 2015. Arsenic and lead uptake by Brassicas grown on an old orchard site. J. Hazard. Mater. 299, 656 663. Available from: https://doi.org/10.1016/ J.JHAZMAT.2015.07.082. Liu, S.-H., Zeng, G.-M., Niu, Q.-Y., Liu, Y., Zhou, L., Jiang, L.-H., et al., 2017. Bioremediation mechanisms of combined pollution of PAHs and heavy metals by bacteria and fungi: a mini review. Bioresour. Technol. 224, 25 33. Available from: https://doi.org/10.1016/J.BIORTECH.2016.11.095. Mileski, G.J., Bumpus, J.A., Jurek, M.A., Aust, S.D., 1988. Biodegradation of pentachlorophenol by the white rot fungus Phanerochaete chrysosporium. Appl. Environ. Microbiol. 54, 2885 2889. Nanseu-Njiki, C.-P., Alonzo, V., Bartak, D., Ngameni, E., Darchen, A., 2007. Electrolytic arsenic removal for recycling of washing solutions in a remediation process of CCAtreated wood. Sci. Total. Environ. 384, 48 54. Available from: https://doi.org/10.1016/ j.scitotenv.2007.04.043. Ojuederie, O.B., Babalola, O.O., 2017. Microbial and plant-assisted bioremediation of heavy metal polluted environments: a review. Int. J. Environ. Res. Public Health 14. Available from: https://doi.org/10.3390/ijerph14121504. Preilipper, U.E.M., Dalfovo, W.C.T., Zapparoli, I.D., Maroubo, L.A., Mainardes, E.L., 2016. Aproveitamento do res´ıduo madeireiro na produc¸a˜o de energia termoele´trica no munic´ıpio de Marcelaˆndia-MT. Desenvolv. e Meio Ambiente 36, 411 428. Available from: https://doi.org/10.5380/dma.v36i0.39802. Punshon, T., Jackson, B.P., Meharg, A.A., Warczack, T., Scheckel, K., Lou Guerinot, M., 2017. Understanding arsenic dynamics in agronomic systems to predict and prevent uptake by crop plants. Sci. Total Environ. 581 582, 209 220. Available from: https:// doi.org/10.1016/J.SCITOTENV.2016.12.111. Rahman, F.A., Allan, D.L., Rosen, C.J., Sadowsky, M.J., 2004. Arsenic availability from chromated copper arsenate (CCA) treated wood. J. Environ. Qual. 33, 173. Available from: https://doi.org/10.2134/jeq2004.1730. Sierra-Alvarez, R., 2009. Removal of copper, chromium and arsenic from preservativetreated wood by chemical extraction-fungal bioleaching. Waste Manag. 29, 1885 1891. Available from: https://doi.org/10.1016/j.wasman.2008.12.015. Townsend, T., Solo-Gabriele, H., Tolaymat, T., Stook, K., Hosein, N., 2003. Chromium, Copper, and Arsenic concentrations in soil underneath CCA-treated wood structures. Soil Sediment Contam. 12, 779 798. Available from: https://doi.org/10.1080/ 10588330390254829. Velizarova, E., Ribeiro, A.B., Ottosen, L.M., 2002. A comparative study on Cu, Cr and As removal from CCA-treated wood waste by dialytic and electrodialytic processes.

References

J. Hazard. Mater. 94, 147 160. Available from: https://doi.org/10.1016/S0304-3894 (02)00063-8. Vidal, J.M., Evangelista, W.V., de, J., Silva, C., Jankowsky, I.P., 2015. Preservac¸a˜o de madeiras no Brasil: histo´rico, cena´rio atual e tendeˆncias. Cieˆncia Florest. 25, 257 271. Vogeler, I., Green, S., Greven, M., Robinson, B., van den Dijssel, C., Clothier, B., 2005. Environmental Risk Assessment of CCA Leaching From Treated Vineyard Posts. HortResearch Client Report No. 17659, 44 p. Wang, L., Chen, S.S., Tsang, D.C.W., Poon, C.-S., Shih, K., 2016. Recycling contaminated wood into eco-friendly particleboard using green cement and carbon dioxide curing. J. Clean. Prod. 137, 861 870. Available from: https://doi.org/10.1016/j. jclepro.2016.07.180. Xie, Y., Sha, Z., Yu, M., 2008. Remote sensing imagery in vegetation mapping: a review. J. Plant. Ecol. 1, 9 23. Available from: https://doi.org/10.1093/jpe/rtm005.

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10

An overview on degradation kinetics of organic dyes by photocatalysis using nanostructured electrocatalyst

Rishu Katwal1, Richa Kothari2 and Deepak Pathania2,3 1

Department of Chemistry, CSKHPKV, Palampur, India Department of Environmental Sciences, Central University of Jammu, Bagla (Rahya-Suchani), Samba, Jammu & Kashmir, India 3 Department of Chemistry, Sardar Vallabhbhai Patel Cluster University, Mandi, Himachal Pradesh, India

2

10.1 Introduction The Earth’s surface covers about 70% of water but only 0.002% of the water is available for human consumption (Alrumman et al., 2016). In the last few decades, due to increases in urbanization and industrialization, the increase in use of heavy metals and dyes in various industries has been observed (Olama et al., 2018). Water pollution occurs when water bodies are adversely affected due to the addition of large amounts of toxic materials. Water pollution is a serious threat to the environment. The organic dyes are major pollutant constituents released into the environment and contaminate soil and water (Bhargava and Jahan, 2012; Das and Charumathi, 2012). A 1971 United Nations report defined ocean pollution as (Oceanic Exploration and Research): The introduction by man, directly or indirectly, of substances or energy into the marine environment (including estuaries) resulting in such deleterious effects to living resources, human health, hindrance to marine activities, including fishing, impairment of quality for use of sea water and reduction of amenities.

Population explosion, extensive industrialization, and advancement in agricultural techniques have been resulted in the release of nondegradable and hazardous material into environment. The textile industry, an important water consumer, produces highly colored and complex wastewater. It has been observed that more than 100,000 commercial dyes with over 7.105 tons of dyestuff were produced yearly. Dyes include a broad spectrum of different chemical structures (Kant, 2012; Lellis et al., 2019). The discharge of dye-containing effluents is undesirable Delivering Low-Carbon Biofuels with Bioproduct Recovery. DOI: https://doi.org/10.1016/B978-0-12-821841-9.00005-0 © 2021 Elsevier Inc. All rights reserved.

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as the dyes breakdown to toxic, carcinogenic, and mutagenic products. Dyes remain in water for a long time. Thus, there is an urgent need for the removal of the various organic dyes from the water system.

10.2 Organic dyes Coloring materials from the past have been used for different purpose such as leather, cloth, food, pottery, and housing. The largest consumption of dyes have been noticed in the textile industry. The residual dyes from different sources such as textile industries, paper and pulp industries, dye-intermediates industries, pharmaceutical industries, tannery and kraft bleaching industries, etc. have been introduced into natural water resources. They may undergo degradation to form products that are highly toxic and carcinogenic. Hence, it is imperative to protect our environment from such toxic pollutants (Lyon: International Agency for Research on Cancer, 2010).

10.3 Classification of organic dyes Organic dyes are classified into two types: 1. Natural dyes Natural dyes are derived from plants and invertebrates. The majorities of natural dyes are vegetable dyes and obtained from plant parts such as roots, berries, bark, leaves, and wood. Natural dyes are often negatively charged. The positively charged dyes are rarely exist. Each dye has been named according to their specific color. The most commonly used natural dye is saffron (natural yellow 6) obtained from the stigmata of Crocus sativus. It is used as an acid dye for coloring and spice of food. Yellow dyes are extracted from cow urine and probably secretions of shellfish and mollusks. Madder or alizarins are another important ancient natural dye reported (Cardon, 2007). 2. Synthetic dyes The dyes derived from organic or inorganic compound are known as synthetic dyes. Due to the presence of heavy metals and complexity in aromatic structure these dyes enhance the mutagenic or carcinogenic toxicity (Mcgeorge et al., 1985). Methods such as biological oxygen demand (BOD), chemical oxygen demand (COD), and total oxygen demand (TOD) detected low ratio value of persistent organic pollutants in wastewater (Wichern et al., 2018; Arin, 1974).

10.4 Methods for the removal of pollutants Efficient techniques for the removal of toxic organic compounds from water have drawn significant interest. A better purification technology reduces the problem

10.5 Advanced oxidation processes

of water shortages, health, energy, and climate change. Due to better solubility in water, the dyes are considered as the common water pollutants. Organic dyes, if present beyond a certain limit, affect the quality of water. So it is necessary to eliminate the dyes from wastewater before being discharged to the environment. A number of methods have been used for the removal of organic and inorganic dyes from wastewater. Some of these are discussed as follows. 1. Biological method The biological process is an economical method for wastewater treatment. In this method, colloidal and dissolved solid is converted into solids by microorganism under favorable environmental conditions. Fungal decolonization, microbial degradation, and bioremediation are commonly used biological methods for the degradation of pollutants (Sharma and Sanghi, 2012). It has been reported that the biological treatment has flexibility in design but requires a large land area. The biological treatment has not been proven to be satisfactory for color elimination. 2. Chemical method In this method, chemicals are used during wastewater treatment to expedite disinfection. Chemical methods include a number of processes such as coagulation or flocculation combined with flotation and filtration, precipitation-flocculation, electrokinetic coagulation, chemical oxidation methods, irradiation, and electroflotation or electrochemical processes. The chemical techniques have been reported for the removal of dyes from the wastewater (Sharma and Sanghi, 2012). 3. Physical methods Physical methods used for the removal of pollutants using natural forces such as gravity, electrical attraction, and Van der Waals forces, etc. Various physical methods such as membrane filtration processes (Reverse osmosis, electrodialysis, nanofiltration) and various adsorption techniques have been employed (Dabrowski, 2001). Among adsorbents, activated carbon has been proven to be the most effective due to its high specific surface area, ultrahigh adsorption capacity, and low selectivity for nonionic and ionic dyes. However, it has some limitations, such as the need for regeneration after exhausting, high cost of the activated carbon, and lack of adsorption efficiency after regeneration (Gong et al., 2009).

10.5 Advanced oxidation processes Advanced oxidation processes are very efficient methods for the generation of strong oxidation species hydroxyl radical;
OH for degradation/mineralization of organic pollutants. The sources of generation of hydroxyl radial (
OH) during the advanced oxidation processes as follows (Naddeo et al., 2015; Chavez et al., 2016; Villegas et al., 2017; Hui et al., 2019):

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Fenton process: Fe21 1 H2 O2 -Fe31 1
OH1 2 OH


Pollutants 1 OH-Degradation products

(10.1) (10.2)

Photo-Fenton process: Fe31 1 H2 O 1 hv-Fe21 1
OH 1 H1


Pollutants 1 OH-CO2 1 H2 O ðMineralizationÞ

(10.3) (10.4)

Photoozonation: O3 1 hv-O2 1 O






H2 O 1 O -H2 O2 -2 OH


Pollutants 1 OH-Degradation products

(10.5) (10.6) (10.7)

Ozonation sonolysis: O3 1 Ultrasonic irradiation-O2 1 O






(10.8)

H2 O 1 O -H2 O2 -2 OH

(10.9)

Pollutants 1
OH-Degradation products

(10.10)

Hydrogen peroxide: H2 O2 1 hv-2
OH

(10.11)

Pollutants 1
OH-Degradation products

(10.12)

10.6 Photocatalysis The word “photocatalytic” is a Greek word and composed of two parts: the prefix “photo” (phos means light) and the word “catalysis” (katalyo means break apart, decompose). The word “photocatalysis” first appeared in a scientific publication in 1911 (Bruner et al., 1911). Later, ZnO reported the reduction of Ag1 to Ag under irradiation. TiO2 and Nb2O5 were used for the reduction of AgNO3 to Ag and AuCl3 to Au (Baur and Perret, 1924; Fujishima et al., 2008). Afterwards, in 1938 TiO2 was also reported as a photosensitizer to bleach dyes in the presence of O2 (Doodeve and Kitchener, 1938). In 1970, Fujishima and Honda introduced photocatalysis in water splitting and later known as the Honda-Fujishima effect (Fujishima and Honda, 1972). One typical example of photocatalysis is the watersplitting reaction described simply as: H2 O 1 Light energy-1/ 2 O2 1 H2

(10.13)

10.6 Photocatalysis

In which the interactions between TiO2 nanoparticles and sunlight split water molecules into oxygen and hydrogen atoms, producing hydrogen gas. In the photocatalysis process, light is used to activate a substance without involving itself in the chemical transformation. The photocatalytic activity depends on the ability of the catalyst to create electron hole pairs, which generate free radicals (e.g., super oxides, hydroxyl radicals) able to undergo secondary reactions (Scheme 10.1). Photocatalysis results in mineralization of organic pollutants, disinfection of water and air, and production of renewable fuels, etc. (Curri et al., 2003). Based on the state of the photocatalyst it can be classified into two categories as: 1. Homogeneous photocatalysis In homogeneous photocatalysis, the reactants and photocatalysts exist in the same phase. Ozone and photo-Fenton systems (Fe1 and Fe1/H2O2) are the examples of homogeneous photocatalysis process. The reactive species in homogeneous catalysis is OH. The mechanism of hydroxyl radical production by ozone can follow two paths (Wu and Chang, 2006). O3 1 hv-O2 1 Oð1DÞ U

U

(10.14)

Oð1DÞ 1 H2 O- OH 1 OH

(10.15)

Oð1DÞ 1 H2 O-H2 O2

(10.16)

H2 O2 1 hv-U OH 1 U OH

(10.17)

However, homogeneous photocatalysis are associated with some major disadvantages. First, the catalysts are stable only in relatively mild conditions, which limit their applicability. Second, the catalysts are dispersed in the phase as the reactant, products, and solvents. The separation at the end of the process is difficult and expensive. In many cases, it is not possible to recover the catalyst. 2. Heterogeneous photocatalysis In heterogeneous reactions, the reactant, product, and catalyst are present in different phases. The transition metal oxides and semiconductors are used as the most common heterogeneous photocatalysts. When a semiconductor

SCHEME 10.1 Photocatalysis process.

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absorbs energy in the form of the photon greater than the band gap of the material, then the electron is excited from the valence band to the conduction band. It results in the generation of positive hole in the valence band. The valence band hole is strongly oxidizing and the conduction band electron is strongly reducing. At the external surface, the excited electron and hole can take part in redox reactions with adsorbed species such as water, hydroxide ion (OH2), organic compounds, or oxygen. Oxidation of water or OH2 by the hole produces the hydroxyl radical (OH), an extremely powerful oxidant. The reduction of adsorbed O2 to OU2 occurs by the conduction band electron. Thus the electron recombining with the hole results in an accumulation of oxygen radical species that can also participate in attacking contaminants (Huang et al., 2017a,b). Heterogeneous photocatalysis has been emerged as an important degradation technology for the mineralization of organic pollutants (Huang et al., 2014). The advantage of the heterogeneous photocatalysis process over other conventional methods is summarized as follows: 1. The processes can be carried out under ambient condition of temperature and pressure. 2. The process uses atmospheric oxygen as oxidant and no other expensive oxidizing chemical is required. 3. The oxidant is strong and less selective, which leads to complete mineralization of organic pollutants. 4. This process is known as green technology. 5. The photocatalysts are cheap, nonhazardous, stable, biologically and chemically inert, and reusable.

10.7 Photocatalysts Photocatalysts are the material which utilizes solar energy for the degradation or mineralization of pollutants. The photocatalysis process is easily available and nontoxic. A promising photocatalyst has properties to absorb a wide range of the solar spectrum of the desired band gap and dissociate the water molecules efficiently. For the past few decades, various metal oxide semiconducting-based nanomaterials and transition have been developed and designed as efficient photocatalysts for watertreatment applications under sunlight. In recent years, many researchers have been working on the modification of photocatalysts. Moreover, various modifications on the photocatalytic material such as heterojunction formation, tuning the morphology, and doping of metals/nonmetals and decorating of nanoparticles onto the photocatalytic surface for the efficient removal of organic dyes and deactivation of microorganism in wastewater under visible light irradiation have been investigated (Zhang et al., 2013; Ge et al., 2019; Chen et al., 2017).

10.8 Photocatalyst surface modifications

10.8 Photocatalyst surface modifications Photocatalyst surface modification increases the charge separation and lifetime of hole for reduction of the recombination process to improve the efficiency of photocatalysis (Karouei and Moghaddam, 2019; Trochowski et al., 2019; Koe et al., 2020). There are three important and efficient modification methods as follows (Scheme 10.2): 1. Metal-semiconductor modification photocatalysts The introduction of metal into the semiconductor transfers the electron from valence band to conduction band of the photocatalyst with narrowing the Eg value and change in electron transfer kinetic with large dipole moment (Chiu et al., 2019). The low Eg value indicates better absorption ability in the visible light or natural sunlight. The value work function of metals determined due to the compact atomic arrangement in Miller indices (111) is most stable. 2. Surface sensitizer photocatalyst In the surface sensitizer process, the physical or chemical adsorption of colored materials absorbs the visible or solar light and excites either to the singlet or triplet excited state. 3. Composite photocatalyst Composite photocatalysts are produced when the big band gap and energy of irradiated light is not enough to irradiate the semiconductor photocatalyst. The other semiconductor with small band gap are coupled to increase the efficiency

SCHEME 10.2 Photoexcitation in photocatalysts: (A) Metal-modification semiconductor; (B) Dye molecule sensitizer; and (C) Composite.

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under near UV or visible and solar light. The photogenerated electrons move from more negative to less negative Fermi energy in conduction band (CB), to prevent charge recombination, while holes flow from more positive to less positive Fermi energy in valance band (VB) (Chiu et al., 2019). However, the composite material performance depends upon microstructure, size, morphology, and adsorption behavior on different surfaces of the photocatalyst.

10.9 Kinetics of photocatalytic degradation The degradation reactions follow a Langmuir-Hinshelwood (L-H) model, the removal rate depends upon a surface area coverage of substrate and work according to the following assumptions (Fox and Dulay, 1993; Mills et al., 1993; Kumar and Chowdhury, 2018): 1. At equilibrium, number of surface site is fixed. 2. At each surface site, only one substrate may connect. 3. For each site, heat of adsorption is the same and independent of surface exposure. 4. No interaction between adjacent adsorbed molecules. 5. The substrate surface adsorption rate is greater than the rate of succeeding chemical reactions. 6. During the binding of the product, no irreversible blocking of active site occurs. r1 5

2 dC kr Kads C 5 dt 1 1 Kads C

(10.18)

where, C is the initial concentration of pollutant, rt is the rate of reaction (changes with time), kr is there action rate constant (mg/L/min), and Kads is the equilibrium or Langmuir adsorption constant of the pollutant molecule on the catalyst surface (L/mg). The initial rate of reaction as a function of C0 can be given by: r1

2 dC k1 Kads C0 5 dt 1 1 Kabc C0

(10.19)

On integrating Eq. (10.18) between the limits: C 5 C0 at time, t 5 0 and C 5 C at t 5 t, we will get: ln

C0 1 Kads ðC0 2 C Þ 5 k1 Kads t C

(10.20)

If Kads C{1, then integrating Eq. (10.18) between the limits: C 5 C0 at t 5 0 and C 5 C at t 5 t, L-H expression can be reduced to 2ln

C 5 kapp t C0

(10.21)

10.10 Photocatalytic reaction parameters

When the organics concentration is low, an “apparent” first-order rate constant Eq. (10.21) could be expressed where (apparent rate constant, kapp 5 k1Kads). Kinetics of water-treatment processes photosensitized by the photocatalyst is dependent on many parameters.

10.10 Photocatalytic reaction parameters 1. Amount/mass of catalyst The amount of catalyst is directly proportional to the rate of reaction. The active sites are increased with the amount of catalysts. This promotes the creation of a greater number of reactive radicals in the photodegradation process. However, above a certain level of catalyst, the reaction rate becomes flat, and at very high amounts of catalyst the penetration of light depresses that introduce to scattering effect (Akpan and Hameed, 2009). 2. Concentration of substrate The concentration of the dye used in the photocatalytic reaction considered one of the most important parameters. The catalyst should be able to degrade an average quantity of the dye. The degradation percentage decreases with increasing quantity of dye concentration (Reza et al., 2017). The amount of the dye is adsorbed or degraded on the surface of the illuminated catalyst, fit with the L-H kinetics model, which assumed that adsorption-desorption kinetics is faster than the photochemical reaction: r5

dC kKC 5 dt ð1 1 KCÞ

(10.22)

where, r is the rate of reaction, k is the rate constants, C is concentration of substrate and t is time of illumination, and K is the Langmuir constant reflecting the adsorption/desorption equilibrium between the substrate and the photocatalyst surface. At low concentrations, the aforementioned equation becomes a first-order equation while at large concentrations of substrate the rate of reaction is maximum and of the zero-order. 3. Initial pH of solution The pH of solution also enhances or suppresses the photodegradation of the organic pollutants. The surface potential of the catalyst varies by changing the pH of the solution, and thereby, three possible mechanism for understanding the effect of pH on the rate of reaction occurs such as hydroxyl radical attack, direct oxidation (positive photohole in valance band of photocatalyst), and direct reduction (photoelectron in the conductive band of the photocatalyst). 4. Temperature In the photocatalytic process, the room temperature is enough to active the photoreaction, while the minimum amount of energy required to promote the

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electron from the valance band to the conductive band of the photocatalyst is called activation energy. The activation energy is calculated by Van ’t Hoff-Arrhenius plot: lnK

2 Ea 1 lnA RT

(10.23)

where, K is the rate constant, Ea is activation energy, T is a temperature of reaction, R is gas constant, and A is a preexponential (frequency) factor. At low temperatures (lower than 0 C), the activation energy increases. While at temperatures above 80 C, the activation energy shows a negative value and depresses the activity of the photocatalytic reaction (Kazuhito et al., 2005; Mamba et al., 2014). 5. Intensity of light Intensity of light is one of the major factors affecting the degradation rate of organic pollutants. At low light intensities the photodegradation rate of reaction is high, and the formation of electron-hole is predominant with negligible recombination process of the electron-hole. At middle light intensities the photodegradation rate of reaction increased directly with an increase of square root of light intensity. At high light intensities, photodegradation rate is independent of light intensity.

10.11 Photocatalytic activity of nonmetals and metalloids supported nanophotocatalyst With environmental pollution a worldwide threat to public health, new initiatives have been developed for environmental restoration for both economic and ecological causes (Bhargava and Jahan, 2012). The contaminated water destroys aquatic life and reduces its reproductive ability. In the United States, nearly 34.8 million tons of hazardous waste was generated in 2005, mostly in the form of liquid waste. Removal of organic pollutants from wastewater is of great concern for researchers due to their harmful effects. Many semiconductors such as TiO2, ZnO, SnO, NiO, Cu2O, Fe3O4, and CdS show great photocatalytic activity under solar light (Julkapli et al., 2014). ZnO NPs have been used effectively for the degradation of several organic pollutants in the presence of sunlight (Pitchaimuthu et al., 2015). ZnO NPs was reported as a photocatalyst for the degradation of methyl orange dye under sunlight (Apollo et al., 2014). Konstantinou and Albanis reported the degradation of azo dyes under sunlight irradiation with the generation of hydroxyl radicals. Hydroxyl radicals are nonselective, strong oxidizers, and very reactive. The free radicals are able to oxidize organic compounds in solution (Konstantinou and Albanis, 2004).

10.11 Photocatalytic activity of nonmetals and metalloids

Another study by reported the incorporation of cerium ions and H2O2 on TiO2, increased the removal of dye in the textile wastewater from 27% to 40% while the color removal increased from around 30% to 55% after 24 hours (Poole and Owens, 2003; Touati et al., 2015). ZnO nanoparticles were reported for the photocatalytic degradation of the environmental pollutants (Preethi, 2018). Pitchaimuthu et al. studied the photocatalytic activity of ZnO powder and compared with TiO2 (Degussa P25) over Acid Brown 14 as the model pollutant (Pitchaimuthu et al., 2015). The excellent photocatalytic activity of TiO2 films in solar light for the degradation of malachite green has been investigated (Sires et al., 2006). ZnWO4 nanoparticle was used for the degradation of rhodamine B in water and decomposition of formaldehyde (Huang et al., 2007). In addition, Lin and Zhu explored ZnWO4 as photocatalysts for the photodegradation of rhodamine B and gaseous formaldehyde (Lin and Zhu, 2007; Akkari et al., 2017). Copper oxide (CuO/Cu2O) is a popular photosensitizer, and photoelectrodes with a large Eg photocatalyst was used to increase the visible light absorption capability for degradation of pollutants (Isherwood, 2016). CuO nanoparticle was reported as an excellent photocatalyst for the degradation of different organic dyes [methylene blue (MB), methyl red (MR), and congo red (CR)] under the illumination of sunlight irradiation. The rate constant for MB, MR, and CR was found to be first-order with values 0.02059, 0.02046, and 0.01749 min21, respectively (Katwal et al., 2015). CuO nanoflower was utilized for photocatalytic activity of MB. The photocatalytic activity of CuO films was determined for the degradation of rose bengal dye (Wang et al., 2009). Cu2O nanocubes for the degradation of dye brilliant red X-3B under simulated solar light (Ma et al., 2010). Al2O3 nanoparticles were explored for photocatalytic degradation of malachite green (MG) dye under sunlight irradiation via two processes: adsorption followed by photocatalysis; and coupled adsorption and photocatalysis. The coupled process exhibited a higher photodegradation efficiency (45%) compared to adsorption followed by photocatalysis (32%) (Pathania et al., 2016). The photocatalytic activity of nanoparticles is affected by the fast recombination of the photogenerate charge carriers which reduces the efficiency of the photocatalytic processes. In order to overcome this problem, simultaneous doping with two kinds of atoms or coupling metal with semiconductor of suitable electronic properties has been investigated (Krishnakumar et al., 2012). It has been found viable strategy to increase the charge separation of the photogenerated electron/hole pairs. Thus, many coupled semiconductor systems have been used as photocatalysts such as ZnOTiO2, ZnOCdS, ZnOAgBr, ZnOAg2S (Subash et al., 2012a,b). Fe-TiO2/rGO (ferum-titanium oxide supported on reduced graphene oxide) photocatalyst was investigated for removal of RhB. The presence of anions such as nitrate, phosphate, chloride, and sulfate the removal efficiency was noted 91%73%, 78%, 48%, and 57%, respectively. It has been observed that COD decreased from and 1550 to 634 mg/L (59.1%) and total organic carbon reduced

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from 930 to 310 mg/L (66.6%) after 390 minutes, showing the photocatalytic efficiently reducing the pollutant in the wastewater (Isari et al., 2018). N-doped TiO2 (NTiO2) and NTiO2 supported on different amounts of RGO (GNTiO2) for photodegradation of organic dye and phenol under visible light irradiation (Zhang et al., 2018). The improved charge separation of the photogenerated electron-hole pairs, due to the heterojunctions between the two oxides, increase the degradation efficiency of nanocomposites (Hamrounia et al., 2014).

10.12 Photocatalytic activity of polymer supported nanophotocatalyst Poly(3,4-ethylenedioxythiophene)/zinc oxide nanocomposites have been used for the photocatalytic efficiency under both UV light than natural sunlight irradiation. The high photocatalytic efficiency under UV light (98.7%) and natural sunlight (96.6%) after 5 hours (Abdiryim et al., 2014) were noticed. A polyaniline-based nanocomposite was reported for photodegradation of MB under solar light (Khan et al., 2014). Graphene-SnO2-PMMA nanocomposite was utilized as photocatalyst to degrade the MB dye under sunlight irradiations with 99% of degradation (Shanmugam et al., 2015). A large number of nonconventional bioadsorbents such as biopolymers have been employed for the remove of toxic metals and dyes from the water system (Zhao et al., 2012; Constantin et al., 2013). Ussia et al. reported nanocomposites combination of 2-hydroxyethyl methacrylate monomer (HEMA), graphene oxide (GO) filler, and ZnO nanolayers for the degradation of MB dye under UV light irradiation. The significant adsorption affinity for MB dye, concentration at 0.73, 0.97, 0.82, and 0.94 mg/g of materials, for pHEMA, pHEMA-GO, ZnO/pHEMA, and ZnO/pHEMA-GO, respectively. This study suggested that pHEMA and pHEMA-GO samples adsorb B37% and B57% of the dye, while ZnO/pHEMA and ZnO/pHEMA-GO samples absorb B41% of MB (Ussia et al., 2018). Very recently, ZnO/PANI nanocomposites were investigated for the degradation of MNZ under UV and visible light irradiation. The degradation rate of MNZ by the ZnO/PANI nanocomposite (k 5 2.53 3 1022 min21) was almost 63 times higher than for the pure ZnO photocatalyst (k 5 0.04 3 1022 min21) (Asgari et al., 2019). PMMA/ ZnO nanocomposite were evaluated for photocatalytic degradation of MB dye under UV light irradiation. About 60% of MB degradation was reported after 4 hours of irradiation (k 5 0.41 3 1022 min21), while compared with ZnO film, degraded nearly 30% of MB after 240 minutes with a photodegradation rate k 5 0.17 3 1022 min21 (Di et al., 2017). Rani and Shanker reported a comparative study on PMMA matrix incorporation of different metal oxides-poly(methyl methacrylate) (MO-PMMA)

References

nanocomposite tested as a photocatalyst for MB degradation. The results obtained for ZnO-PMMA, Ni2O3-PMMA, CuO-PMMA, and Fe3O4-PMMA shows degradation percentage of 99%, 98%, 93%, and 90%, respectively (Rani and Shanker, 2018). Pectin-based CuS nanocomposites were reported for photocatalytic degradation of MB under sunlight irradiated with about 95% degradation. TiO2/ZnO/chitosan was evaluated for the photocatalytic decolorization of methyl orange in aqueous solution under solar light 97% of degradation was observed within 4 hours (Gupta et al., 2012, 2013). Gupta et al. reported pectin-based nanocomposite for photocatalytic degradation of MB and MG dyes. 89.21% and 79.27% of degradation were reported within 3 hours of photo irradiation using MB and MG, respectivly. Pathania et al. studied pectin-based nanocomposite for the photocatalytic degradation of methylene blue dye in the presence of solar irradiation. It was recorded that 97.02% of MB dye is degraded after 60 minutes of irradiation (Gupta et al., 2012, 2014, 2015). Most recently, Xylan/polyvinyl alcohol (PVA)/TiO2 composite were reported with excellent photodegradation of ethyl violet and astrazon brilliant red 4G dyes under visible irradiation. The studies revealed the degradation rates of ethyl violet and astrazon brilliant red 4G dyes were 93.65% and 92.71%, respectively after 60 minutes in presence of visible irradiation.

10.13 Conclusions The literature data indicate a high efficiency of photocatalytic process in the oxidative degradation of water pollutants. The visible-light-responsive photocatalyst treatment will be utilized for large-scale commercial applications. The unique features of photocatalysts with a view of their complete mineralization have been briefly discussed in this chapter. The photocatalytic efficiency is not only influenced by the nature of the photocatalysts, but also affected by the different reaction parameters were evaluated. Thus, the photocatalysis provide advanced efficiency and good robustness for wastewater treatment.

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Reza, K.M., Kurny, A.S.W., Gulshan, F., 2017. Parameters affecting the photocatalytic degradation of dyes using TiO2: a review. Appl. Water Sci. 7 (4), 15691578. Available from: https://doi.org/10.1007/s13201-015-0367-y. Shanmugam, M., Alsalme, A., Alghamdi, A., Jayavel, R., 2015. Photocatalytic properties of graphene-SnO2-PMMA nanocomposite in the degradation of methylene blue under direct sunlight irradiation. Mater. Express 5 (4), 319325. Available from: https://doi. org/10.1166/mex.2015.1246. Sharma, S.K., Sanghi, R., 2012. Advances in Water Treatment and Pollution Prevention. Springer Science and Business Media, XX, 460, https://doi.org/10.1007/978-94-0074204-8. Sires, I., Garrido, J.A., Rodr´ıguez, R.M., Cabot, P.L., Centellas, F., Arias, C., et al., 2006. Electrochemical degradation of paracetamol from water by catalytic action Fe21, Cu21, and UVA light electrogenerated hydrogen peroxide. J. Electrochem. Soc. 153, D1D9. Available from: https://doi.org/10.1149/1.2130568. Subash, B., Krishnakumar, B., Pandiyan, V., Swaminathan, M., Shanthi, M., 2012a. An efficient nanostructured Ag2SZnO for degradation of acid black 1 dye under day light illumination. Sep. Purif. Technol. 96, 204213. Available from: https://doi.org/ 10.1016/j.seppur.2012.06.002. Subash, B., Krishnakumar, B., Velmurugan, R., Swaminathan, M., Shanthi, M., 2012b. Synthesis of Ce co-doped AgZnO photocatalyst with excellent performance for NBB dye degradation under natural sunlight illumination. Catal. Sci. Technol. 2, 23192326. Available from: https://doi.org/10.1039/c2cy20254a. Touati, A., Hammedi, T., Najjar, W., Sayadi, S., 2015. Photocatalytic degradation of textile wastewater in presence of hydrogen peroxide: effect of cerium doping Titania. J. Ind. Eng. Chem. 35. Available from: https://doi.org/10.1016/j.jiec.2015.12.008. Trochowski, M., Kobielusz, M., Mro´z, K., Surowka, M., Hamalainen, J., Iivonen, T., et al., 2019. How insignificant modifications of photocatalysts can significantly change their photocatalytic activity. J. Mater. Chem. A 43. Available from: https://doi.org/10.1039/ C9TA09400H. Ussia, M., Di, Mauro, A., Mecca, T., Cunsolo, F., Cerruti, P., Nicotra, G., et al., 2018. ZnO-pHEMA nanocomposites: an eco-friendly and reusable material for water remediation. ACS Appl. Mater. Interfaces 10, 4010040110. Available from: https://doi.org/ 10.1021/acsami.8b13029. Villegas, P., Stefanos, G., Rtimib, G.S., Grandjeanc, D., Bensimonc, M., Felippede, L., et al., 2017. A green solar photo-Fenton process for the elimination of bacteria and micropollutants in municipal wastewater treatment using mineral iron and natural organic acids. Appl. Cata. B: Environ. 219, 538549. Available from: https://doi.org/ 10.1016/j.apcatb.2017.07.066. Wang, S., Xu, H., Qian, L., Wang, J., Liu, Y., Tang, W., 2009. CTAB assisted synthesis and photocatalytic property of CuO hollow microshperes. J. Solid. State Chem. 182, 10881093. Available from: https://doi.org/10.1016/j.jssc.2009.01.042. Wichern, M., Gehring, T., Lu¨bken, M., 2018. Reference module in earth systems and environmental sciences. Model. Biol. Syst. Treatise Water Sci. 4, 231263. Wu, C.H., Chang, C.L., 2006. Decolorization of reactive red 2 by advanced oxidation processes: comparative studies of homogeneous and heterogeneous systems. J. Hazard. Mater. 128 (23), 265272. Available from: https://doi.org/10.1016/ j.jhazmat.2005.08.013.

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213

Index Note: Page numbers followed by “f” and “t” refer to figures and tables, respectively.

A ABR. See Anaerobic batch reactor (ABR) Acetate, 35 36, 123 124 Acetate production, 152 153 Acetic acid production, 134 135 Acetogenesis process, 35 Acetogenic microbes, 152 153 Acidogenesis process, 35 Activated-carbon air-cathodes, 5 6, 8 Activation energy, 203 204 Active aeration, 59 61 Adsorption-desorption kinetics, 203 Advanced oxidation processes, 197 198 AEM. See Anion exchange membrane (AEM) Aerobic process, 52 55 Agro wastewater, 36 Air-cathode MFC, 59 61 Algae, 55 56 Algal biofilm MFC, 55 56 Algal bioreactor, 70 71 Algal cathode MFC, 55 56 Al2O3 nanoparticles, 205 Ammonia, 45 48, 50 51 in cathode chamber, 61 62 diffusion of, 59 61 recovery, 59 62 solidifying, 59 Ammonium, 52 migration, 50 51 nitrogen, 48 and phosphate, 61 62 Ammonium bicarbonate, 59 61 Ammonium hydroxide, 153 Ammonium ion, electromigration of, 59 61 Ammonium sulfate, 59 61 Anaerobic ammonium oxidation (ANAMMOX), 45 46, 48 49, 51 52 bacteria, 51 52 Anaerobic anode chamber, 107 108 Anaerobic batch reactor (ABR), 38 39 Anaerobic biodegradation, 92 Anaerobic bioprocesses, 166 Anaerobic digestion, 41 Anion exchange membrane (AEM), 52 55 Anodic biofilm, 10, 12f Anodic chamber, 169 ANOVA analysis, 167 Artificial wastewater (AW), 9

Autotrophic bacteria, 48 Autotrophic nitrification, 48 Azo dyes remediation, 90

B Bacillus licheniformis, 181 182 Bacteriological approaches, 58 59 BES. See Bioelectrochemical systems (BESs) Betaproteobacteria, 58 59 Biocathode, 91, 127 128, 130 Biochemical oxygen demand (BOD), 101 102, 107, 119 120 removal, 115, 115f Biocommodities production, 155t Bioelectricity, 59 61 Bioelectrochemical systems (BESs), 85, 124, 143 144 ammonia recovery, 59 62 ammonia removal in, 50 51, 50f anode electrode in, 124 anodic chamber of, 166 bacteriological approaches, 58 59 biohythane generation in, 166 167, 169 171 denitrifying biocathode, 58 59 development in, 62 66 electro-biocommodities generation in, 154 156 electron transfer and chemical synthesis in, 125f groundwater remediation using, 56 57 implementation, 143 144 influential operational parameters, 57 58 interdependency with, 145f nitrogen. See Nitrogen phosphorus removal and recovery, 66 74, 67f, 72t. See also Phosphorus principle of, 86 role in remediation of pollutants, 90 93 scaling up of technology, 94 sustainability of technology, 93 94 theory, 86 types of enzymatic fuel cells for energy production, 88 89 microbial desalination cells, for energy production, 90 microbial electrolysis cells for energy, 88 microbial electrosynthesis for energy production, 88 microbial fuel cells, 87 88

215

216

Index

Bioelectrochemical systems (BESs) (Continued) microbial solar cells for energy production, 89 plant microbial fuel cells, 89 working mechanism of, 144 146 Bioenergy production enzymatic fuel cells for, 88 89 microbial desalination cells for, 90 microbial electrolysis cells for, 88 microbial electrosynthesis for, 88 microbial solar cells for, 89 plant microbial fuel cells for, 89 Bioethanol production, 151 152 Biohydrogen production, 32, 149, 165, 173t Biohythane, 165 166 generation in bioelectrochemical system, 166 167, 169 171 production, 168f reactor types for, 170t substrate for, 167 169 technology, 173 use of, 172 Biological method, 197 Biological nitrification, 49 Biological nitrogen removal (BNR) processes, 48 energy-efficient alternative to, 48 Biomethane production, 149 151 Bioremediation, 90 91, 180 183 Blocking oscillator, 24f BOD. See Biochemical oxygen demand (BOD) Brown rot fungi, 184 Butyric acid, 153

C Calcium phosphate, 71 74 “The Californian Energy Crisis”, 2 Candidatus Brocadia sinica, 51 52 Candidatus Kuenenia, 51 52 CANON process. See Completely autotrophic nitrogen removal over nitrite (CANON) process Carbamate, 61 62 Carbon brush anode, 8 Carbon dioxide, 152 153 Cathode catalysts, 38 Cathode chamber, 61 62 Cathode reduction potential, 33 Cathodic chamber, 166 Cationic exchange membrane (CEM), 38 39, 46 CBD. See Controlled biocathodic denitrification systems (CBD) CCA-treated wood poles, 179f, 188t CEM. See Cationic exchange membrane (CEM)

Channa striatus (haruan), 109 110 Chemical methods, 197 Chemical oxygen demand (COD), 35, 101 102, 107, 119 120 removal, 115 116, 116f Chemical remediation, 186 189, 187f Chlorella vulgaris, 55 56 Chloronitrobenzene remediation, 91 Chlorophyta, 89 Chronopotentiometry, 6 Clarias batrachus, 109 110, 119 Clostridium carboxidivorans/Clostridium ragsdalei, 127 Clostridium ljungdahlii, 126 Clostridium tyrobutyricum BAS7, 153 C/N ratio, 57 58 COD. See Chemical oxygen demand (COD) COD-containing wastes, 167 169 Completely autotrophic nitrogen removal over nitrite (CANON) process, 48 49 Composite photocatalyst, 201 202 Coniophora puteana, 180 181 Continuous vs. intermittent mode of operation, 18 20, 19f Controlled biocathodic denitrification systems (CBD), 56 57 Conventional denitrification, 51 Conventional nitrogen removal process, 47 48 Copper oxide, 205 Coulombic efficiency (CE), 39 40 Coupled aerobic anoxic nitrous decomposition operation (CANDO) process, 48 49 Crystallization process, 39 CuO nanoparticle, 205

D Dark fermentation (DF), 32, 35 DC/DC converter, 21 Degradation photocatalytic, 202 203 of pollutants, 197, 200 Denitrification, 49, 51 55 Denitrifying bacteria (Thauera), 58 59 Desulfobulbus propionicus, 152 153 Digital potentiometer, 6 Digital-to-analog converter (DAC), 6 2,4-Dintrochlorobenzene, 91 Direct connection, 144 Direct electron transfer mechanism, 34 35 Dissimilatory nitrate reduction to ammonium (DNRA), 58 59 Dissolved oxygen (DO), 39 40 Diverse microbial species, 87

Index

DNRA. See Dissimilatory nitrate reduction to ammonium (DNRA) DO concentration, 57 58 Domestic waste water, 36 38 Dual-chamber BES, 47f, 56 57, 172f Dual-chamber MFC, 59 61 Dye-containing effluents, 195 196

E EAB. See Electroactive bacteria (EAB) EBPR. See Enhanced biological phosphorus removal (EBPR) Economic evaluation, 134 136 Economic viability, 93 94 Electricity, 101 Electroactive bacteria (EAB), 58 59 Electroactive microbes, 123 Electro-active microbial community, 40 Electroactive microorganisms, 90 Electro-biocommodities, 148 153 Electrochemically active bacteria (EAB), 32, 33f, 143 144, 147 diverse groups of, 34 35 Electrochemically active biofilm, 57 58 Electrodes, 5 6, 111 113, 114f disposition, 5 6 materials, 128 129 potential effect, 129 130 Electrolyte, 4 5 Electromigration of ammonium ion, 59 61 Electromotive force, 6 Electron acceptor, 51 52 Electron donor, 51 Electron transfer, 87 Electron transfer mechanism (ETM), 34 35 Electrosynthesis products, 137 Energy conversion process, 3 intensive process, 31 32 module structure diagram, 23f Energy production, 21 25 process, 4 Energy regulation and storage, 21 25 field effect transistor driver, 24 operation stage, 22 23 oscillator, 23 starting stage, 22 voltage comparator, 23 24 voltage supervisor, 25 Energy/water expenditure data, 1 2 Energy-water nexus, 1 concept of, 2

Enhanced biological phosphorus removal (EBPR), 66 67 Environmental pollution, 204 Environmental risks of treated wood, 178 180 Enzymatic fuel cells, for energy production, 88 89 Escherichia coli, 58 59 Ethylenediamine tetraacetic acid (EDTA), 187f ETM. See Electron transfer mechanism (ETM) European directive 2012/27/EU, 2 3 Exchange membranes, 86 Exoelectrogenic anaerobic bacteria, 4 5 Exoelectrogens/electrogens, 34 35, 58 59 Exoeletrogens (Geobacter), 58 59 External resistance, 57 58 Extracellular electron transfer, 58 59 Extracellular polymeric substance (EPS), 154 156 Extracted energy values, 10, 11f, 14f, 15f, 16f

F Faraday’s constant, 33 Fenton process, 198 Fermentation effluent, 38 39 Fertilizer production process, 45 Field effect transistor (FET) driver, 24, 25f Fish-processing wastewater characteristics, 105 107, 105t, 112f physiochemical parameters, 105 107 biochemical oxygen demand, 107 chemical oxygen demand, 107 nitrogen and phosphorus, 107 odor, 106 organic content, 106 pH, 105 106 solids content, 106 temperature, 106 as substrate, 109 110 using microbial fuel cell, 110 113 electrode, 111 113 physical, chemical, and biological parameters, 110, 112t substrate, 110 Fluctuations in electricity supply, 126 127 Fomitopsis palustris, 180 181 Forward osmosis (FO) system, 59 61 Fossil fuels (FFs), 31 32 Fuel Diversification Policy, 102 103 Functional elements, 4 Fungi, 180 in remediation process, 183 186

217

218

Index

G Gammaproteobacteria, 58 59 Gas diffusion cathode, 61 62 Geobacter, 58 59 Geobacteraceae, 87 Geobacter lovleyi, 58 59 Geobacter metallireducens, 51 Geobacter species, 51 Geobacter sulfurreducens, 58 59 Gibbs free energy, 34 of reactions, 134 135 Gluconic acid, 184 185 Glycolysis pyruvate, 184 185 Graphene-SnO2-PMMA nanocomposite, 206 Graphite electrodes, 87 88 Green technology, 103 Groundwater remediation by ex situ approach, 56 57 using bioelectrochemical system, 56 57

H Haber Bosch process, 47 48, 59 Halocarbon cis-dichloroethene, 91 Heavy metals, 184 removals in MEC, 39 40 treatment, 93 Heterogeneous photocatalysis, 199 200 Heterotrophic denitrification, 56 58 Homogeneous photocatalysis, 199 Honda-Fujishima effect, 198 199 HRT. See Hydraulic retention time (HRT) Human population, 1 Hybrid-graphite felt, 166 Hybrid MDC MEC, 39 Hydraulic retention time (HRT), 51 52 Hydrogen, 149 Hydrogen peroxide, 198 production, 153 Hydrogen production, 31 32, 41 MEC technology for, 35 36, 37t Hydrogen yield of pyrolysis, 36 Hydroxyapatite, 68 Hydroxyl radicals, 204

I IEEE1451 standard, 3 4, 27 Implemented control strategy, 22 Incorporated BESs-AD system, 149 151 Indirect contact, 144 Indirect electron transfer mechanism, 34 35 Industrial carbon dioxide emission, 143 Industrial wastewater, 38

Influential operational parameters, 57 58 Inoculum, impact of, 126 127 Inorganic pollutants treatment, 92 93 Integrated MEC approach, 40 41 Integrated MEC-MFC system, 51 52 Integrated pyrolysis, 36 Intensity of light, 204 Interconnected anodes, 8 Intermittent aeration, 52 55 Interrupted load operation, 18 20, 19f Ion exchange membrane (IEM), 59 61 ISO14046, 2 3

J Junction gate field-effect transistor (JFET), 24

L Lactic acid producing bacteria, 182 Lactobacillus acidophilus, 182 Lactobacillus bulgaricus, 182 Lactobacillus plantarum, 182 Laetiporus sulphureus, 180 181 Langmuir-Hinshelwood (L-H) model, 202 203 LCA. See Life cycle assessments (LCA) Legislation, 2 3 Life cycle assessments (LCA), 2 3 Lignocellulose biomass compounds, 36 renewable characteristics of, 36 LTC2935 device, 25

M Magnesium ammonium phosphate hexahydrate, 68 Magnesium hydroxide, 68 Magnesium phosphate, 68 Malaysia Plan for energy development, 104t Market saturation, 135 136 MEC. See Microbial electrochemical systems (MEC) MEDC. See Microbial electrolysis desalination cell (MEDC) Mediated ETM, 34 35 Metal-semiconductor modification photocatalysts, 201 Methane, 165 Methanobacterium palustre, 149 Methanogenesis, 41 Methanol-rich wastewater, 38 Methyl viologen (MV), 151 152 MFC. See Microbial fuel cells (MFCs) Microalgae (Chlorella vulgaris), 45 46, 70 71 Microbes, 87 Microbial-based electrochemical process, 86

Index

Microbial community analysis, 51 52 Microbial desalination cells, for energy production, 90 Microbial electrochemical systems (MEC), 32, 123 124 agro wastewater, 36 domestic waste water, 36 38 extracellular electron transport during, 34f fermentation effluent, 38 39 fundamentals and working principles, 32 34 industrial wastewater, 38 integrated approach, 40 41 nutrient and heavy metals removals in, 39 40 principal components of, 32 technology in hydrogen production using wastewater, 35 36, 37t two-chambered, 33f Microbial electrochemical technology (MET), 39 in wastewater treatment, 146 147, 148f Microbial electrolysis cells (MECs), 39, 46, 166 for energy, 88 Microbial electrolysis desalination cell (MEDC), 59 61 Microbial electrosynthesis cell (MES), 86, 123, 137, 144 145, 148, 150f acetate production in, 129t basic principle of, 124 economic evaluation, 134 136 electrode materials, 128 129 electrode potential effect, 129 130 for energy production, 88 factors affecting, 124 131 fluctuations in electricity supply, 126 127 future scope of work, 136 137 impact of inoculum, 127 128 microbial community on, 126 127 pH effect, 125 126 reactor design effect, 130 131 reactors used in, 132t strategies to improve product titer, 131 134 Microbial fuel cells (MFCs), 3 21, 46, 87 88, 101, 144 145 algal cathode, 55 56 basic components of, 108t batch operating mode, 3 with cathode composition, 5f closing remarks, 20 21 continuous vs. intermittent mode of operation, 18 20 energy extraction, 6 21 for energy production, 17 18, 21 25 fish-processing wastewater characteristics, 105 107, 105t improving energy production, 8 20

internal impedance, 3 National Green Technology Policy, 103 107 operation efficiency, 3, 23 24 performance of, 49 polarization curve on, 10 regulation and storage, 21 25 series and parallel association, 12 18, 13t smart sensor structure and operation, 25 27 substrates used in, 109 system, 107 110 theoretical analysis, 4 6 treatment methodology of fish-waste using, 110 113 used for power improvement, 9t waste from fresh markets, 103 105 Microbial nutrient recovery cell (MNRC), 70 71 Microbial solar cells for energy production, 89 Microcontroller unit (MCU) activation, 22, 26 Microorganisms, 180 MNRC. See Microbial nutrient recovery cell (MNRC) Monosaccharide, 152 Multidisciplinary system, 143 144 Multiple-cycle method, 7 9

N Nafion, 107 108 Nanophotocatalyst, 204 206 National Green Technology Policy (NGTP), 103 107 Natural dyes, 196 Nematoloma spp., 180 Nernst equation, 6, 33 Network interconnection, 27 NGTP. See National Green Technology Policy (NGTP) Nitrate-containing groundwater, 56 57 Nitrate/nitrite, 51 as electron acceptors, 51 reduction process, 52 55 toxicity, 45 Nitrobenzene compounds remediation, 91 Nitrogen, 59, 107, 116 118 conversion process, 47 48 one-way ANOVA analysis for, 117t, 118t removal and recovery, 47 66, 63t in bioelectrochemical system, 49 59, 53f challenges in, 62 66 issues related to conventional technologies, 48 49 transformation, 52 56 Nitrogen removal efficiency, 39 40 Nitrophenol compounds, 91

219

220

Index

Nutrients in MEC, 39 40 removal and recovery of, 45

O Oil-rich fish, 109 110 OLR. See Organic loading rate (OLR) Omega-3 fatty acids, 109 110, 109t One-way Analysis of Variance, 113, 117 119, 117t, 118t, 119t Oreochromis niloticus (tilapia), 109 110 Organic-anaerobic processes, 41 Organic compounds, 52 57 Organic dyes, 196 197 advanced oxidation processes, 197 198 classification of, 196 kinetics of photocatalytic degradation, 202 203 photocatalysts, 200 photocatalyst surface modifications, 201 202 pollutants removal, methods for, 196 197 Organic loading rate (OLR), 36 38 Organic matter removal efficiency, 38 Organic pollutants, 49 Organic products, 153 Oscillator, 23 Oxalate, 184 Oxalic acid, 181, 184 Oxidation-reduction reaction, 6 Oxidative wastewater, 146 147 Ozonation sonolysis, 198 Ozone and photo-Fenton systems, 199

P Palm oil mill effluent (POME), 40 41 Pangasius sutchi, 109 110, 119 PAO. See Phosphate accumulating organism (PAO) PBESs. See Photo-bioelectrochemical systems (PBESs) Pennisetum setaceum, 89 Perchlorate ions, 92 93 Periodic sampling process, 26 Petrochemical compounds, 92 pH, 105 106 effect of, 125 126 PHA. See Polyhydroxyalkanonate (PHA) Phanerochaete chrysosporium, 180, 185 Phosphate accumulating organism (PAO), 66 67 Phosphate rocks, 66 Phosphorus, 45 46, 66, 107, 118 119, 118f issues related to, 66 67 for polyphosphate production, 66 67 removal and recovery, 66 74, 72t

challenges in, 71 74 in single-chamber BES, 71 74 struvite precipitation, 68 Photo-bioelectrochemical systems (PBESs), 55 56 Photocatalysis process, 198 200, 199f Photocatalysts, 200 photoexcitation in, 201f surface modifications, 201 202 Photocatalytic activity of nanoparticles, 205 of nonmetals and metalloids, 204 206 of polymer supported nanophotocatalyst, 206 207 Photocatalytic degradation, kinetics of, 202 203 Photocatalytic reaction parameters, 203 204 Photo-Fenton process, 198 Photo-MFC, 55 56 Photoozonation, 198 Physical methods, 197 Physiochemical parameters, 105 107 Pilot plant application, 62 66 Planar cathode, 5 6 Planktomycetes, 48 Plant microbial fuel cells for energy production, 89 Polarization curve, 7 9, 7f, 12 for big volume reactor, 12f of small reactor, 14f, 16f Pollutants removal, 196 197 Polyaromatic hydrocarbons, 92 Polychlorobiphenyl pollutants, 91 92 Polyhydroxyalkanonate (PHA), 66 67 Polyhydroxybutyrate, 152 Polymer supported nanophotocatalyst, 206 207 Polysaccharides, 167 169 Poly(3,4-ethylenedioxythiophene)/zinc oxide nanocomposites, 206 Potentiostat, 6 7 Product chamber, 61 62 Protons, 124 Pseudomonas aeruginosa, 92

R Ralstonia eutropha, 152 RBC. See Rotating biological contractor (RBC) Reactor design effect, 130 131 Reference voltage, 22 24 Remediation process of bromate and chlorate, 92 93 chemical, 186 189, 187f heavy metals treatment, 93 inorganic pollutants treatment, 92 93 mechanisms used by fungi in, 183 186

Index

of organic xenobiotics, 90 92 azo dyes remediation, 90 chloronitrobenzene remediation, 91 nitrobenzene compounds remediation, 91 polyaromatic hydrocarbons, 92 of polychlorobiphenyl pollutants, 91 92 and recovery of treated wood, 180 189 bioremediation, 180 183 Renewable bioenergy, 101 Renewable sources of energy, 85 Rhodopseudomonas palustris, 51 52 Rotating biological contractor (RBC), 52 55 Rot fungi, 181

S Sediment-type PBES, 70 71 Sensor network, 4 Series and parallel association, 12 18, 13t SHARON process, 48 49 Shewanella, 87 Shewanella oneidensis, 58 59, 90, 154 156 Shewanella putrefaciens, 58 59 Simultaneous nitrification and denitrification (SND), 52 55 Single-chamber MFC, 107 108 Single-chamber reactors, 5 6 Single-cycle method, 7 8 Smart sensor general overview of, 26f structure and operation, 25 27 Smart transducer, 25 26 SND. See Simultaneous nitrification and denitrification (SND) Soil contamination, 180 Solids content in wastewater, 106 Spartina angilica, 89 Spirulina, 55 56 Sporomusa ovata, 127, 148, 152 153 SSM. See Stainless steel mesh (SSM) Stainless steel mesh (SSM), 39 Standard Gibbs free energy of formation, 9, 10t State-of-the-art approach, 59 61 Streptococcus thermophilus, 182 Struvite, 45 46 Struvite precipitation, 68 Substrates for biohythane generation, 167 169 concentration of, 203 fish-processing waste as, 109 110 in microbial fuel cell, 109 preparing, 110 Subsurface water, nitrate toxicity of, 45 Surface sensitizer photocatalyst, 201

Synthetic dyes, 196 Synthetic wastewater, 52 56, 59 61

T TCE. See Trichloroethene (TCE) Temperatures, 203 204 Textile industry, 195 196 Thauera spp., 58 59 Thermochemical process, 31 32 3D RVC electrodes, 128 Trametes versicolor, 185 Treated wood CCA in, 188, 188t chemical remediation, 186 189, 187f environmental risks of, 178 180 remediation process of, 186 Trichloroethene (TCE), 91 TSAF. See Two-step anaerobic fermentation (TSAF) Two-chamber MFC, 108f Two-phase DF, 36 Two-stage anaerobic digestion, 41 Two-step anaerobic fermentation (TSAF), 173

U Ugold, 61 62 Uninterrupted operation mode, 18 U-Power, 70 71 Urine, 61 62, 70 71

V Value-added biochemical’s production, 148 153 acetate production, 152 153 bioethanol production, 151 152 biohydrogen production, 149 biomethane production, 149 151 hydrogen peroxide production, 153 other, 153 Van ’t Hoff-Arrhenius plot, 203 204 VC. See Voltage comparator (VC) VFAs. See Volatile fatty acids (VFAs) Volatile fatty acids (VFAs), 66 67, 125 Voltage comparator (VC), 22 24, 24f Voltage production, 113 114, 114f Voltage supervisor, 25

W Waste-to-wealth concept, 35 Wastewater, 52 55 development, 4 oxidation of, 144

221

222

Index

Wastewater treatment, 101 102 bioelectricity generation with, 147 MEC technology for, 35 36, 37t microbial electrochemical technology in, 146 147 Wastewater treatment plants (WWTP), 3, 20 21, 66 67 Water-energy nexus, 27 Water footprint assessment, 2 3 Water pollution, 195 Water scarcity, 1 2

Water treatment, 2 White rot fungi, 185 Wood preservatives, 178, 178t Wood vinegar, 188 189 WWTP. See Wastewater treatment plants (WWTP)

Z ZnO nanoparticles, 204 205 ZnWO4 nanoparticle, 205