Current Developments in Biotechnology and Bioengineering: Biochar Towards Sustainable Environment 0323918735, 9780323918732

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Current Developments in Biotechnology and Bioengineering

Series Editor: Professor Ashok Pandey Centre for Innovation and Translational Research CSIR-Indian Institute of Toxicology Research Lucknow, India & Sustainability Cluster School of Engineering University of Petroleum and Energy Studies Dehradun, India

Current Developments in Biotechnology and Bioengineering Biochar Towards Sustainable Environment Editors Huu Hao Ngo Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia

Wenshan Guo Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia

Ashok Pandey Centre for Innovation and Translational Research, CSIR-Indian Institute of Toxicology Research, Lucknow, India; Sustainability Cluster, School of Engineering, University of Petroleum and Energy Studies, Dehradun, India

Sunita Varjani Gujarat Pollution Control Board, Gandhinagar, Gujarat, India

Daniel C.W. Tsang Department of Civil and Environmental Engineering, The Hong Kong Polytechnic University, Kowloon, Hong Kong, China

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 © 2023 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. ISBN: 978-0-323-91873-2 For information on all Elsevier publications visit our website at https://www.elsevier.com/books-and-journals

Publisher: Susan Dennis Editorial Project Manager: Helena Beauchamp Production Project Manager: Kiruthika Govindaraju Cover Designer: Miles Hitchen Typeset by STRAIVE, India

Contributors Mukesh Kumar Awasthi College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, PR China Sanjeev Kumar Awasthi College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, PR China Yazid Bindar Department of Chemical Engineering; Department of Bioenergy and Chemurgy, Faculty of Industrial Technology, Institut Teknologi Bandung, Bandung, Indonesia Yogi Wibisono Budhi Department of Chemical Engineering, Faculty of Industrial Technology, Institut Teknologi Bandung, Bandung, Indonesia S. Woong Chang Department of Environmental Energy Engineering, Kyonggi University, Suwon, Republic of Korea Zhuo Chen Environmental Simulation and Pollution Control State Key Joint Laboratory, State Environmental Protection Key Laboratory of Microorganism Application and Risk Control (SMARC), Beijing Laboratory for Environmental Frontier Technologies, School of Environment, Tsinghua University, Beijing, PR China Dongle Cheng Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Lijuan Deng Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Quoc Cuong Do Chemical & Process Technology Division, Korea Research Institute of Chemical Technology (KRICT); Department of Civil and Environmental Engineering, Korea Advanced Institute of Science and Technology (KAIST), Daejeon, Republic of Korea Vivek K. Gaur Amity Institute of Biotechnology, Amity University Uttar Pradesh, Lucknow, India Wenshan Guo Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Zizhang Guo Shandong Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Technology, Shandong University, Qingdao, China

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Contributors

Mingjing He Department of Civil and Environmental Engineering, The Hong Kong Polytechnic University, Kowloon, Hong Kong, China Pandit Hernowo Department of Chemical Engineering, Faculty of Industrial Technology, Institut Teknologi Bandung, Bandung, Indonesia Zhen Hu Shandong Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Technology, Shandong University, Qingdao, China Wei Jiang School of Environmental Science and Engineering, Huazhong University of Science and Technology, Wuhan, PR China Jianxiong Kang School of Environmental Science and Engineering, Huazhong University of Science and Technology, Wuhan, PR China Yan Kang College of Environment and Safety Engineering, Qingdao University of Science and Technology, Qingdao, China Gajasinghe Arachchige Ganga Kavindi Faculty of Life and Environmental Sciences, University of Tsukuba, Ibaraki, Japan Sunil Kumar CSIR-National Environmental Engineering Research Institute, Nagpur, India D. Duong La Institute of Chemistry and Materials, Hanoi, Vietnam Zhongfang Lei Faculty of Life and Environmental Sciences, University of Tsukuba, Ibaraki, Japan €t Dresden, Dresden, Huanyu Li Institute of Construction Materials, Technische Universita Germany; School of Naval Architecture, Ocean and Civil Engineering, Shanghai Jiao Tong University, Shanghai, China Shuang Liang Shandong Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Technology, Shandong University, Qingdao, China Dongqi Liu School of Environmental Science and Engineering, Huazhong University of Science and Technology, Wuhan, PR China Misha Liu Faculty of Life and Environmental Sciences, University of Tsukuba, Ibaraki, Japan Tao Liu College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, PR China Xiaoning Liu State Key Laboratory of Water Resources and Hydropower Engineering Science, School of Water Resources and Hydropower Engineering, Wuhan University, Wuhan, PR China

Contributors

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Yang Liu Joint Research Centre for Protective Infrastructure Technology and Environmental Green Bioprocess, School of Environmental and Municipal Engineering, Tianjin Chengjian University, Tianjin, China Yi Liu Department of Environmental Science and Engineering, Fudan University, Shanghai, China €t Viktor Mechtcherine Institute of Construction Materials, Technische Universita Dresden, Dresden, Germany Amin Mojiri Department of Civil and Environmental Engineering, Graduate School of Advanced Science and Engineering, Hiroshima University, Hiroshima, Japan Huu Hao Ngo Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Dinh Duc Nguyen Department of Environmental Energy Engineering, Kyonggi University, Suwon, Republic of Korea Manh Khai Nguyen Faculty of Environmental Sciences, University of Science, Vietnam National University, Hanoi, Vietnam Thu Thuy Nguyen Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Bing-Jie Ni Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Ashok Pandey Centre for Innovation and Translational Research, CSIR-Indian Institute of Toxicology Research, Lucknow; Sustainability Cluster, School of Engineering, University of Petroleum and Energy Studies, Dehradun, India Ashutosh Kumar Pandey Centre for Energy and Environmental Sustainability, Lucknow; CSIR-National Environmental Engineering Research Institute, Nagpur, India Anping Peng Joint Research Centre for Protective Infrastructure Technology and Environmental Green Bioprocess, School of Environmental and Municipal Engineering, Tianjin Chengjian University, Tianjin, China Yongzheng Ren School of Environmental Science and Engineering, Huazhong University of Science and Technology, Wuhan, PR China Syed Saquib Department of Chemical Engineering, Faculty of Industrial Technology, Institut Teknologi Bandung, Bandung, Indonesia Tjandra Setiadi Department of Chemical Engineering, Faculty of Industrial Technology, Institut Teknologi Bandung, Bandung, Indonesia

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Contributors

Xingdong Shi Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Yuqing Sun Department of Civil and Environmental Engineering, The Hong Kong Polytechnic University, Kowloon, Hong Kong, China B.X. Thanh Key Laboratory of Advanced Waste Treatment Technology, Vietnam National University Ho Chi Minh (VNU-HCM), Ho Chi Minh City, Vietnam Thi Hien Tran Institute of Environmental Science, Engineering and Management, Industrial University of Ho Chi Minh City, Ho Chi Minh City, Vietnam Thi Nhung Tran Department of Civil and Environmental Engineering, Korea Advanced Institute of Science and Technology (KAIST), Daejeon, Republic of Korea Van Son Tran Faculty of Environmental Sciences, University of Science, Vietnam National University, Hanoi, Vietnam Daniel C.W. Tsang State Key Laboratory of Clean Energy Utilization, Zhejiang University, Hangzhou; Department of Civil and Environmental Engineering, The Hong Kong Polytechnic University, Kowloon, Hong Kong, China Sunita Varjani Gujarat Pollution Control Board, Gandhinagar, Gujarat, India Steven Wahyu Department of Chemical Engineering, Faculty of Industrial Technology, Institut Teknologi Bandung, Bandung, Indonesia Dan Wang Integrated Research of Energy, Environment and Society (IREES), Energy and Sustainability Research Institute (ESRIG), University of Groningen, Groningen, The Netherlands Lei Wang State Key Laboratory of Clean Energy Utilization, Zhejiang University, Hangzhou, China Wei Wei Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Jonathan W.C. Wong Institute of Bioresource and Agriculture, Hong Kong Baptist University, Kowloon Tong, Hong Kong Lan Wu Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Huijun Xie Environmental Research Institute, Shandong University, Qingdao, China Jingtao Xu School of Municipal and Environmental Engineering, Shandong Jianzhu University, Jinan, China

Contributors

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Jian Yang School of Naval Architecture, Ocean and Civil Engineering, Shanghai Jiao Tong University, Shanghai, China Yuanyao Ye School of Environmental Science and Engineering, Huazhong University of Science and Technology, Wuhan, PR China Jian Zhang College of Safety and Environmental Engineering, Shandong University of Science and Technology; Shandong Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Technology, Shandong University, Qingdao, China Xinbo Zhang Joint Research Centre for Protective Infrastructure Technology and Environmental Green Bioprocess, School of Environmental and Municipal Engineering, Tianjin Chengjian University, Tianjin, China Yuying Zhang Department of Civil and Environmental Engineering, The Hong Kong Polytechnic University, Kowloon, Hong Kong, China Zengqiang Zhang College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, PR China John Zhou School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Yuwen Zhou College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, PR China

Preface The book titled Biochar Towards Sustainable Environment is a part of the Elsevier comprehensive book series on Current Developments in Biotechnology and Bioengineering (Editor-in-Chief: Ashok Pandey). Biochar, as a renewable material, can be produced from various sustainable biomass feedstocks through pyrolysis technologies. Reuse of biomass wastes for biochar production is a sustainable strategy for biowaste management and environmental protection. The obtained biochar with specific physicochemical and surface characteristics can be further applied in water and wastewater purification, construction and drainage systems, soil remediation, sustainable agriculture development, resource recycling, energy storage and conversion, as well as climate change mitigation. This book highlights the contribution of biochar to environmental sustainability. The book provides a detailed overview of the sustainable biomass wastes feedstocks and different technologies for biochar production, and its sustainable applications in various fields. The book comprises 16 chapters. Chapter 1 discusses the sustainable biochar production technologies including sustainable biomass resources for biochar, reviews of available biochar production technologies, projections on biochar demand, environmentally friendly and intensified biochar production technologies, and trade and economy of biochar. Chapter 2 describes the reuse of various biowastes as feedstocks for biochar production sustainably in biowaste management while focusing on the environmental benefits of this strategy. Chapter 3 deals with the tailored production of engineered biochar, roles and interactions of biochar in construction products, significance of chemical compositions and physical properties, and mechanical performance and functionality of biochar-augmented construction products. Chapter 4 focuses on the significance and design of sustainable drainage systems, identified needs for biochar amendment, physical improvement by biochar amendment, chemical improvement by biochar amendment, biological improvement by biochar amendment, and future research directions. Chapter 5 introduces the advances in the sustainable application of biochar for water purification, application of different biochar in water environment, mechanism of biochar in water purification, and sustainable application examples. Chapter 6 describes the application of biochar for treating wastewater not only as the main treatment method but also as the pretreatment and the posttreatment methods in integrated treatment systems. Chapter 7 summarizes the preparation and physicochemical characteristics of biochar and introduces the performance of nutrients (i.e., nitrogen and phosphorus) recovery from wastewater. This chapter also discusses the existing challenges, future research efforts, and opportunities for biochar in nutrients recovery from wastewater. Chapter 8 outlines xix

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Preface

comprehensively the recent progresses and breakthroughs in the use of biochar in the field of sludge treatment including sludge dewatering, aerobic/anaerobic sludge digestion, and anaerobic sludge fermentation by applying its specific properties. This chapter provides a theoretical basis and technical reference for the application of biochar for improving sludge treatment. Chapter 9 reviews storm water reuse, harvesting issues, characterization of storm water, potential for reuse, and the quantity and characteristics of biomass. The modifying techniques for biochar production and the application and future perspectives of biochar for storm water reuse are also discussed. Chapter 10 provides a review of current biochar properties, its use as an adsorbent/ amendment for soil remediation, and its effect on microorganism communities as well as plant growth. In addition, the contribution of biochar to bioeconomy is discussed. Chapter 11 deals with the evolution of biochar production technologies for sustainable agriculture purposes, modification of biochar for sustainable agriculture, influence of biochar on soil nutrient dynamics and enzymes, and impact of biochar on crop growth and yield. Chapter 12 provides an overview of the commercial biochars and its applications for energy storage and conversion. The challenges and opportunities together with future perspectives are also discussed. Chapter 13 describes the application of biochar in polyaromatic hydrocarbons remediation. Integration of biochar with other available technologies has been discussed for environment management. Knowledge gaps and perspectives in polyaromatic hydrocarbons remediation are summarized. Chapter 14 compares the production conditions, physiochemical properties, and existing and potential applications between biochar and hydrochar in addition to future challenges and research directions, particularly from an environmentally sustainable perspective. Chapter 15 provides an overview of the significance, application, and future developments of biochar and introduces the sustainability assessment concept in terms of environmental, economic, social, and integrated aspects for the evaluation of the application of biochar towards sustainability. Chapter 16 focuses on identifying the sustainability impacts of biochar production, in combination with its application via an integrated life-cycle assessment framework that utilizes three methodologies: life-cycle assessment (LCA), life-cycle costing (LCC), and social life-cycle assessment (S-LCA). We express our deepest appreciation to the authors and reviewers for their valuable contributions to the book. We are also very grateful to the Elsevier team comprising Dr. Kostas Marinakis, former Senior Acquisitions Editor; Dr. Katie Hammon, current Senior Acquisitions Editor; Bernadine A. Miralles, Editorial Project Manager; and the entire production team of Elsevier for supporting us constantly during the editorial process. Huu Hao Ngo Wenshan Guo Ashok Pandey Sunita Varjani Daniel C.W. Tsang

1 Sustainable technologies for biochar production Yazid Bindara,b, Yogi Wibisono Budhia, Pandit Hernowoa, Steven Wahyua, Syed Saquiba, and Tjandra Setiadia a DE PARTMENT OF CHEMICAL ENGINEERING, FACULTY O F INDUSTRIAL T ECHNOLOGY, INSTITUT TEKNOLOGI B ANDUNG, BANDUNG, INDONE SIA b DEP ARTME NT OF BI OE NERGY AND CHEMURGY, FACULTY OF INDUSTRIAL T ECHNOLOGY, I NSTITUT TEKNOLOGI BANDUNG, BANDUNG, INDONESIA

1. Introduction Biochar is a solid product leftover from the thermal conversion of biomass. This thermal conversion process is known as the pyrolysis. Biomass pyrolysis gives three products at once. The first product is a noncondensable gas named bio-pyrolysis gas (BPG). The second one is a liquid product resulting from the condensation of volatile products released by the biomass pyrolysis process. This liquid product is commonly termed as Bio-oil. However, it presents in a crude form, i.e., bio-crude oil (BCO) which will further processed for proper utilization. The remaining solid material resulting from the pyrolysis is referred to as Biochar. Biochar has a complex chemical molecular structure with molecular formula which is still not known with certainty. It is chemically quantified from the mass fraction of the leading organic elements that make it up. The main organic elements consist of carbon (C), hydrogen (H), and oxygen (O). Besides them, Biochar is also formed by minor elements, if any, such as nitrogen (N), sulfur (S), chlorine (Cl), and others. It also contains inorganic compounds such as mineral elements that are present in form of ash. Biochar stability is described using proximate analysis. This method provides information on the mass fraction of biochar that thermally decomposed at a temperature of 900°C. It also give information regarding the mass fraction of various other constituents including volatile matter, fixed carbon and ash content. High mass fraction of volatiles indicates higher degradability potential of Biochar. While high fixed carbon content resulted in high stability against the degradation in the soil. Biochar particles as a whole are formed by organic material resulting from the carbonization process, uncarbonized organic material, and inorganic ash formed by various minerals (Ok et al., 2016). Carbonized organic matter biochar has a high C content. This material is formed by carbon bonds with a structure that is still amorphous, without constructing crystalline structure like graphite. Specific functional groups forms the surface of such organic matter, then pores formed inside it. The morphological characteristics of Biochar describe the existence of these holes in a microview. Current Developments in Biotechnology and Bioengineering. https://doi.org/10.1016/B978-0-323-91873-2.00013-3 Copyright © 2023 Elsevier Inc. All rights reserved.

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Biochar production technology schemes are generally based on the heating method, heating rate and raw material feeding. The first scheme involves an indirect heating method to biomass for biochar production with the absence of any combustion reaction with oxygen at temperature above 250°C (Basu, 2013). The second scheme employs a direct heating method by supplying the hot combustion gases to the pyrolysis chamber to undergo the pyrolysis reactions of the biomass. The third scheme is the burning the same biomass on top of the biomass for pyrolysis by supplying very small amount of the air in pyrolysis chamber. In this method, the produced pyrolysis gas is burned together with biomass in the top section of pyrolysis chamber. For a small-scale, this production technology was shown by Shepard (2011). Another biochar production technology is based on continuous technology where the pyrolysis chamber is in the form of a cylinder whose biomass is driven by a screw. This pyrolysis cylinder gets heating indirectly through its outer wall. The biochar product that comes out of the pyrolysis cylinder is fed to the separation chamber for obtaining biochar and the resulting volatile gas. Biochar is then discharged from the bottom of this chamber through a screw conveyor and cooled down using cold water. Such production technology is already exists on a commercial scale known as the Pyreg technology (Pyreg, 2021; Fesharaki and Rath, 2018). Various biochar production technologies were developed with different technological schemes; however, each one has there advantages and disadvantages. Sustainable and environment friendly technology with lesser pollution is an essential requirement for large scale commercial biochar production with high economic output. A continuous biochar production technology with a processing capacity of 240 kg/day biomass produces 72 kg/day of biochar with diesel fuel for heating an auger-type pyrolyzer. The economic selling price of a biochar was set at US$ 1165 per ton with the biomass feed stock cost at US$ 72 per ton while the market price itself was around US$ 2300 per ton (Pawar and Panwar, 2020). A high yield of quality biochar achieved without releasing volatile materials and without direct contact of later with water is the preferred process flow. The biochar production process developed for commercial use should also maintain simpler process technology. Biochar produced from the fast or lightning pyrolysis provides lower yields, more complex production technology, and high production costs. Biochar production technology developed must meet the rules of a high recovery rate, a production capacity that can meet the needs, and attractive economic feasibility. Besides this, the production technology must rely only on the biomass energy source without involving external energy sources to become sustainable. The use of materials other than biomass which makes the production technology unsustainable and causes high production costs should be avoided. The fulfillment of environmentally friendly production technology is a crucial requirement for large scale production. The pyrolysis of biochar releases volatile materials which may cause pollution when discharged into the environment. Volatile materials consisting of hydroxycarbon compounds must be burned in the combustion chamber to only liberate CO2 and H2O which are nonhazardous polluting agents. The problems that exist are indeed related to the supply of biomass that can be provided. Since the available biomass

Chapter 1 • Sustainable technologies for biochar production

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is mostly residue from agricultural products, the availability of this biomass is not centralized. It is provided at certain times of the year in mere quantities. The determined biochar production technology must adapt to such biomass supply conditions.

2. Biomass as raw material for production of biochar The raw material used for the production of biochar is biomass. This biomass includes intentionally grown raw material, residue from crops, residue from forestry products, or organic waste. Biomass that is intentionally grown as raw material generally does not consider as food and has a short growing life. Agricultural residue biomass generally has no or very low economic value, such as rice husks and straw. Biomass obtained from organic waste is disposed of or sorted as a material with no economic value anymore. Sources of agricultural residue biomass include corn (Tippayawong et al., 2018), wheat (Sedmihradska´ et al., 2020), rice (Tsai et al., 2021), barley (Jazini et al., 2017), sorghum € z, 2014), olive (Naik et al., 2017), soybean (Kong et al., 2011), rapeseed (Angin and Senso (Abdelhadi et al., 2017), oil palm (Idris et al., 2014), sunflower (Klimek-Kopyra et al., 2021), coconut (Castilla-Caballero et al., 2020), cassava (Tippayawong et al., 2017), sugarcane (Jeong et al., 2016), beet (Yao et al., 2011), coffee (Kiggundu and Sittamukyoto, 2019), cotton (Shen et al., 2015), and others. The world’s potential for residual biomass in 2017 is theoretically estimated in between 4000 and 12,000 million tons (WBA (Producer), 2019). The estimated residual biomass potential is higher than the 2010 estimate, which was around 3300 million tons (Born et al., 2014). Of this residual biomass potential, 10% is used as feed, 78% is returned to agricultural land in the form of burning or biomass alone (Born et al., 2014). Wood production for fuel was reported as 1944 million m3 for 2019. This wood was converted into wood pellets as much as 38.9 million tons and into charcoal as much as 53.1 million tons for 2019 (WBA (Producer), 2020). Estimates for biomass planted for energy use are also based on the availability of 3500 million ha of land with biomass productivity of 8 tons/ha per year (Hoogwijk et al., 2005). Biomass from plants is divided into wood and nonwood. The primary constituent of wood biomass is cellulose and lignin (Basu, 2013). In addition to lignocellulose, plants also contain extractives and ash. For examples, switchgrass is formed by 40% cellulose, 25% hemicellulose, and 15% lignin (Miles et al., 1995), and fruit empty bunch consists of 44.4% cellulose, 24.3% hemicellulose, and 31.3% lignin (Nasser et al. (2016). Each of lignocellulosic components has a different chemical structure. Biomass has been designated as a renewable energy source. Biomass conversion into thermal energy is done through burning. For this conversion to thermal energy, biomass is quantified by the content of volatile materials as products of thermal decomposition, the fixed carbon content as organic solids left after the biomass is thermally decomposed, and the ash content as materials that do not contain thermal energy. Biomass analysis for this composition is known as proximate analysis, which provides data on the mass fraction of volatile matter (VM), fixed carbon material (FC), ash material (A), and water (M) owned by biomass. Its

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calorific value of combustion measures the potential energy content of biomass, reported in high calorific value (HHV) or low calorific value (LHV). These values are used in the energy assessment of biomass. The current biomass energy assessor is compared to the energy content of coal. The level of its energy content also referred to measures the economic value.

3. Biochar production process technology Biochar production technology is generally based on a thermochemical process. The thermochemical process commonly used in biochar production technology is pyrolysis (Brown, 2009; Mullen et al., 2010; Manya` et al., 2018; Kazawadi et al., 2021). Another thermochemical method used for biochar along with synthesis gas production is the gasification process. Another method used for biochar production along with synthesis gas production is the gasification process (Yao et al., 2018). The hydrothermal process that uses water as a heating medium at high pressure and temperature to convert biomass into biochar products has also received attention in the biochar production. The product of this process is often called hydrochar (Sharma et al., 2019). Pyrolysis methods in biochar production based on the operating conditions are expressed as slow pyrolysis, intermediate pyrolysis, fast pyrolysis, and flash pyrolysis (Table 1). The slow pyrolysis occurs at low heating rate. Pyrolysis with moderate heating rate is categorized as an intermediate pyrolysis. The fast pyrolysis itself takes place at high heating rate. Flash pyrolysis is carried out by heating the pyrolyzed biomass very rapidly in a matter of milliseconds (Basu, 2013). The required time to heat up the pyrolyzed biomass at pyrolysis temperature Tpf is referred as heating time tht. The heating rate β is approached as ΔT/tht and considered to be constant during the process. The total time between the heating time and the pyrolysis reaction time can be defined as the residence time tct of the pyrolyzed biomass. Other parameters to categorize the pyrolysis process condition are the temperature level, residence time and biomass particle diameter. The slow pyrolysis is operated at the heating rate below 60°C/min. The fast pyrolysis employs the heating rate in the range of 600–12,000°C/min, whereas flash pyrolysis use the heating rate above 60,000°C/min (Balat et al., 2009). The intermediate pyrolysis is carried out at heating rate about 200°C/min (Jouhara et al., 2018) or in the range of 60°C/min to 600°C (Tripathi et al., 2016). The yield of biochars becomes lower for higher heating rate at the same Table 1

Pyrolysis category. Pyrolysis method

Operating conditions

Slow pyrolysis

Intermediate pyrolysis

Fast pyrolysis

Flash pyrolysis

Temperature (°C) Heating rate (°C/min) Biomass residence time (s) Biomass particle diameter (mm)

300–600 6–60 >1800 >5

400–650 60–600 600 1–5

850–1250 600–12,000 0.5–10 1000 >60,000 Mg2+ >NH+4 > K+ > Na+. It meansCa2+ and Mg2+ are the main cations for ion exchange on the biochar. Jing et al. (2019) analyzed using XRD characterization that Mg2+ was embedded in the prepared Mg(OH)2/montmorillonite composite biochar, and Mg2+ located in the interlayer space exchanged with NH+4 in water through ion exchange, the maximum adsorption capacity of the composite biochar for NH+4 was up to 50.27 mg/g. Surface complexation Biomass pyrolysis is a process of dihydroxylation/dehydrogenation and aromatization. Various functional groups (dOH, dCOOH, C]O, etc.) are formed on the biochar, and the pyrolysis temperature also affects the type and number of functional groups, which play an important role in the process of nitrogen adsorption by biochar (Yin et al., 2017). Yang et al. (2017) found that the pine sawdust biochar prepared by pyrolysis at 300°C may keep more cellulose chains and lignin functional groups, and oxygencontaining groups on the surface of the biochar (such as dCOOH, C]O) NH+4 bonding, thus increasing the adsorption of NH+4 . FTIR infrared spectroscopy analysis showed that the stretching vibration peaks of functional groups such as dCOOH and C]O have changed, indicating that the polarity and hydrophilicity of the biochar surface are beneficial to

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the bonding of dCOOH and C]O functional groups with NH+4 (Cui et al., 2016). These results confirm the oxygen-containing functional groups play an important role in the adsorption of NH+4 by biochar. Similarly, Hu et al. (2020) prepared three kinds of pericarp biochar and obtained similar results through FTIR infrared spectrum analysis. It was found that after the adsorption of NH+4 in water by pericarp biochar, the stretching vibration peaks of dOH and C]O functional groups were weakened, indicating these functional groups participated in the adsorption of ammonium through surface complexation.

3.2 Phosphorus recovery Excess phosphorus discharge into the water environment could cause serious eutrophication, resulting in the imbalance of the ecosystem and the impairment of sustainable development of humans and society. Therefore, how to efficiently remove excessive phosphorus from water has aroused extensive attention. Meanwhile, phosphorus is widely used and in great demand, but its reserves are limited and nonrenewable. Therefore, more and more studies attempt to focus on the recovery of phosphorus resources from anthropogenic discharged phosphorus. As mentioned before, biochar has become a widely used material for phosphorus removal and recovery from water. Table 4 summarizes some recent researches regarding phosphate removal from water with biochar. However, raw biochar is not conducive to the adsorption of phosphate due to the limited surface area and lack of functional groups, as well as the significant negative charge on the surface. To overcome these limitations, modified biochar adsorbents with metal substances were commonly used in previous studies for adsorption and recovery of phosphorus. Herein, the performance, influencing factors, and mechanism of biochar in adsorbing phosphate from wastewater are briefly summarized.

3.2.1 Performance of biochar on phosphorus adsorption In general, biochar sorption performance on phosphate is largely dependent on the raw materials that use to make biochar, producing procedures, and modified methods. As shown in Table 3, most of the raw materials using to produce biochar are agricultural and forestry wastes, with a small number of animal manure and sludge. The pyrolysis temperature is mostly between 300°C and 600°C and rarely is above 800°C. Metal modification is the most used modification method, which generally achieves highly improved adsorption capacity. The reported adsorption capacity of the modified biochar for phosphate is between 2.628 and 758.96 mg/g. Some research results show that the adsorption of phosphate by biochar follows the Langmuir adsorption isotherm model and the pseudo-second-order kinetic model, indicating that the adsorption of phosphate on biochar is monolayer adsorption. However, some other studies showed that the Langmuir-Freundlich adsorption isotherm model is also suitable to describe the adsorption of phosphate on biochar o adsorbed phosphate. Zhu et al. (2020) reported the adsorption process of phosphate on modified corn stover biochar is fitted the Langmuir-Freundlich model with the R2 value of 0.997, indicating that

Table 4

Adsorption of phosphorus in water by biochar.

Feedstock

Pyrolysis temperature (°C) Modifier

Initial Adsorption concentration capacity (mg/L) (mg/g)

Bamboo

600

Mg-Al/Mg-Fe LDH

5–600

172

Cabbage, Rape

500

Mg-Al LDO

5–500

127.2–132.8

Date palm waste fronds

700

Mg-Al LDH

10–50

177.97

Langmuir

Sugarcane leaves

550

Mg-Al LDH

5–500

81.83

Langmuir

Cattail

400–600

La(OH)3

5–500

36.06

Langmuir

Platanus balls

600

La(OH)3

30–100

148.11

Langmuir

Wheat straw

550

25–300

108.86

Langmuir

Pineapple peels

300

Quaternary chitosan, La(OH)3 Fe2O3, La(OH)3

100–400

101.16

Langmuir

Oak

500

La(OH)3

1–400

46.37

Langmuir

Dewatered sewage sludge

400–800

La(OH)3

50–300

93.91

Langmuir

Corn straws

800

NaLa(CO3)2, FeCl3

50–400

217.84

Langmuir

Kinetic Adsorption Isotherm model mechanism Langmuir Pseudosecondorder Langmuir Pseudosecondorder

Ion exchange

Electrostatic attraction, Ligand exchange, Ion exchange Pseudo- Ligand exchange firstorder Pseudo- Electrostatic second- attraction, Ligand order exchange Pseudo- Ligand exchange, second- Electrostatic order attraction Pseudo- Electrostatic second- attraction, Ligand order exchange Pseudo- Electrostatic second- attraction, Ligand order exchange Pseudo- Surface precipitation, second- Ligand exchange, order Electrostatic attraction Pseudo- Electrostatic second- attraction, Surface order precipitation, Ligand exchange Pseudo- Electrostatic second- attraction, Ligand order exchange Pseudo- Electrostatic second- attraction, Ligand order exchange

References (Wan et al., 2017) (Zhang et al., 2019)

(Alagha et al., 2020) (Li et al., 2016)

(Xu et al., 2019c) ( Jia et al., 2020) (Huang et al., 2020) (Liao et al., 2018a)

(Wang et al., 2016)

(Li et al., 2020b) (Qu et al., 2020)

Continued

Table 4

Adsorption of phosphorus in water by biochar—cont’d

Feedstock

Pyrolysis temperature (°C) Modifier

Initial Adsorption concentration capacity (mg/L) (mg/g)

Kinetic Adsorption Isotherm model mechanism

Lagerstroemia indica leaves

600

Mg(OH)2

10–300

121.95

Sips

Longleaf pine, Red oak, Hard maple 500

MgCl2

100–600

28.2–29.22

Cypress sawdust

400–600

MgCl2

50–250

43.5–66.7

Hickory wood chips, Bamboo

600

MgCl2, AlCl3, FeCl3

15–1000

119.6

Sugarcane harvest residue

550

MgCl2

1–500

398.71

Moso bamboo

400–600

MgCl2

19.7–498

344–370

Banana straw, Cassava straw, 430 Chinese fir straw, Corn straw, Taro straw, Camellia oleifera shell Corn straw 550

MgCl2

20–350

6.77–31.15

MgCl2

4–200

60.95

Carrots

400

MgCl2

25–350

138

Corn cob

800

MgCl2

3–25

8.44

Hickory wood chips

600

AlCl3

–

8.346

Poplar chips

550

AlCl3

50–1600

57.49

AlCl3

0–3000

701.65– 758.96

Poultry manure, Sugarcane straw 350–650

References

Pseudo- Ligand exchange (Luo et al., second2021a) order Freundlich – Surface precipitation (Oginni et al., 2020) Langmuir Pseudo- Surface precipitation (Haddad et al., second2018a) order Langmuir- – Electrostatic (Zheng et al., Freundlich attraction, Surface 2020) precipitation Langmuir Pseudo- Electrostatic (Li et al., second- attraction, Surface 2017b) order precipitation Langmuir- Pseudo- Ligand exchange, (Jiang et al., Freundlich second- Electrostatic 2018) order attraction Langmuir Pseudo- Electrostatic (Jiang et al., second- attraction, Surface 2019) order precipitation Langmuir- Pseudo- Electrostatic (Zhu et al., Freundlich second- attraction, Surface 2020) order precipitation Langmuir Pseudo- – (de Carvalho firstEufra´sio Pinto order et al., 2019) Langmuir Pseudo- – (Jena et al., second2021) order Langmuir Pseudo- – (Zheng et al., second2019) order Langmuir- Pseudo- – (Yin et al., Freundlich second2018b) order Langmuir – – (Novais et al., 2018)

Wheat straw

600

213.22  13.57 Langmuir- PseudoFreundlich secondorder 111.80– Sips Pseudo129.79 secondorder 24.9–27.6 Freundlich Pseudosecondorder 5.068 Langmuir- PseudoFreundlich secondorder 111 Freundlich Pseudosecondorder

CaCl2

200–1000

Peanut shells, Sugarcane bagasse 460–850

MgCl2, CaCl2

3–5800

Rice husk, Wood

500

Fe2(SO4)3, FeSO4

25–150

Water hyacinth

450

FeCl3, FeCl2

0.186–150

Waste-activated sludge

550

FeCl3

5–1000

Corn straw

550

FeCl3

10–50

98

Walnut shell

600

FeCl2, MgCl2

5–400

6.945

Distillers grains

600

Phosphogypsum 5–500

102.4

Oilseed rape straw

700

Coal gangue

5–200

7.9

Sheep manure

800

Oyster shells, LaCl3

5–200

88.34

Surface precipitation (Li et al., 2020c) Surface precipitation (Fang et al., 2020)

Electrostatic attraction, Ligand exchange Electrostatic attraction, Ligand exchange Electrostatic attraction, Ion exchange, Ligand exchange Langmuir – Ion exchange, Surface precipitation, Ligand exchange Langmuir- Pseudo- Electrostatic Freundlich second- attraction order Freundlich Pseudo- Electrostatic second- attraction, Surface order precipitation, Ligand exchange Langmuir Pseudo- Electrostatic second- attraction, Ligand order exchange, Surface precipitation Langmuir Pseudo- Ligand exchange secondorder

(Ajmal et al., 2020b) (Cai et al., 2017) (Yang et al., 2018a)

(Min et al., 2020) (Tao et al., 2020) (Wang et al., 2020)

(Wang et al., 2021)

(Feng et al., 2021)

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the adsorption of phosphate by biochar was controlled by multiple processes. In addition, Fang et al. (2020) and Luo et al. (2021a) found that the adsorption of phosphate with modified peanut shells, bagasse, and crape myrtle leaf biochar is well in line with the Sips adsorption isotherm model. Generally speaking, the adsorption process of phosphate with biochar was relatively complicated.

3.2.2 Influencing factors Temperature The adsorption efficiency of biochar for phosphate generally increases with the increase of temperature (Wang et al., 2016; Xu et al., 2019b). This may attribute to the adsorption process of phosphate by biochar is a spontaneous endothermic reaction with a negative Gibbs free energy (ΔG) and a positive enthalpy (ΔH). Besides, the temperature increase can also accelerate the diffusion rate of phosphate ions to the adsorption site (Xu et al., 2019b). Qu et al. (2020) used magnetic corn stover biochar modified by NaLa(CO3)2 and FeCl3 to adsorb phosphate. The maximum adsorption capacity predicted by Langmuir adsorption isotherm was increased from 253.98 to 330.86 mg/g when the solution temperature was raising from 15°C to 35°C. Interestingly, studies have shown that the adsorption efficiency of biochar for phosphate only increases within a certain temperature range. If the reaction temperature is too high, the adsorption sites on the surface of the biochar may be affected, resulting in a decrease in adsorption capacity. For example, Jia et al. (2020) showed that when the temperature increased from 25°C to 35°C, the maximum adsorption capacity of lanthanum-loaded Platanus balls biochar for phosphate increased from 124.61 to 148.11 mg/g, but as the temperature continued to rise to 40°C, the adsorption capacity dropped to 123.55 mg/g. Initial concentration of phosphate The initial concentration of phosphate in the solution could also significantly affect the adsorption efficiency of biochar. On the one hand, the initial concentration of phosphate would affect the concentration gradient between the biochar surface and the solution by reducing the mass transfer resistance, thus affecting the gradient dynamics and promoting the adsorption of phosphate by biochar (Tao et al., 2020). On the other hand, an increase in the initial concentration of phosphate could also increase the collision rate between the adsorbate and the adsorbent, thereby increasing the adsorption capacity (Haddad et al., 2018b). For example, Luo et al. (2021b) explored the adsorption effect of La(OH)3 modified walnut shell biochar on phosphate can be explained by the intraparticle diffusion model. Due to the large difference in the initial phosphorus concentration and the abundant active sites on the surface of biochar, the fitting curve is divided into two parts, and the adsorption rate in the first stage is greater than that in the second stage. However, with the increase of adsorption time, the adsorption capacity gradually decreases due to the decreases of concentration gradient and active site on biochar.

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215

In addition, the initial concentration of phosphate will also affect the absorption rate and the equilibrium time. For example, Li et al. (2020c) conducted a study on the adsorption kinetics of modified wheat straw biochar to different concentrations of phosphate (100, 200, and 300 mg/L). The adsorption rates at the three concentrations were all faster at the beginning and then reached equilibrium gradually. The difference was that the higher the initial concentration of the phosphate, the longer it took to reach the adsorption equilibrium. The adsorption equilibrium is 20, 240, and 360 min with the initial concentrations of 100, 200, and 300 mg/L, respectively. pH As a key factor, the pH can affect the removal of phosphorus by biochar in three ways. First, phosphorus can form different species at different pH because phosphoric acid has three pKa values (pKa1 ¼ 2.12, pKa2 ¼ 7.21, pKa3 ¼ 12.67). The main phosphorus species are H3PO4, H2PO4
, HPO42
and PO43
at solution pH < 2.12, 2.12 < pH < 7.21, 7.21 < pH < 12.67, and pH > 12.67, respectively. (Fig. 2) (Yin et al., 2018a). The adsorption energy between biochar and four phosphorus species is different, so the removal efficiency of biochar on phosphorus is discrepant. Meanwhile, whether the metal ions on the surface of biochar can form precipitation with different forms of phosphate ions will also affect the adsorption effect. Li et al. (2020c) explored the adsorption effect of nanoCaO biochar on phosphate under different pH conditions (3.04–11.01). At lower pH value,

FIG. 2 Distribution curve of phosphoric acid.

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Current Developments in Biotechnology and Bioengineering


phosphate mainly exists in the form of H2PO
4 , and H2PO4 cannot form a precipitate with 2+ Ca on the surface of biochar. Second, the pH of the solution will affect the charge of the biochar surface. When pH < pHzpc, the surface of biochar is positively charged, and when pH > pHzpc, the surface of biochar is negatively charged (Jiang et al., 2019). Studies have shown that the electrostatic attraction between biochar and phosphate can only be formed when the surface of biochar is positively charged, thereby increasing the adsorption efficiency and promoting the removal of phosphate from water. Conversely, if the surface of the biochar is a negative charge, the adsorption effect will be lowered due to electrostatic repulsion. Li et al. (2020a) studied the influence of pH on the adsorption of phosphate by lanthanum-loaded sludge biochar. The results showed that when pH < 6.0, the adsorption capacity of biochar remained at a high level, and with pH further increased, the adsorption efficiency declined rapidly. The pHzpc of the biochar was 6.99, which was consistent with the above conclusion. Third, when pH is at a high value, there are more hydroxide ions in water, which will compete with phosphate for adsorption sites on biochar, thus reducing the adsorption capacity of biochar.

Coexisting ions There are many kinds of ions in the actual water environment, and the existence of these ions might affect the removal efficiency of phosphate by biochar. The common anions
2

that affect the removal of phosphate by biochar include SO2
4 , NO3 , CO3 , HCO3 and
Cl , etc. These coexisting anions mainly affect the adsorption process in two ways. On the one hand, coexisting anions will compete with phosphate for adsorption sites on biochar. On the other hand, some anions can also affect the adsorption process by changing
the pH values of the solution, such as CO2
3 and HCO3 . Jia et al. (2020) found that increas2

ing the concentration of SO4 and NO3 did not affect the phosphate adsorption process of
lanthanum-modified Platanus biochar, while the addition of CO2
3 and HCO3 could significantly inhibit the phosphate removal efficiency of the biochar, and the inhibition increased with the increase of the concentration of both. The reasons for the inhibition
can be mainly divided into the following two aspects: first, CO2
3 and HCO3 can occupy the adsorption sites on the surface of biochar, and then form La2(CO3)3 and La(HCO3)3 with lanthanum loaded on the surface of biochar. Second, the addition of these two ions will increase the pH of the solution, thus increasing the electrostatic repulsion between biochar and phosphate. Different coexisting anions have different effects on the removal of phosphate by bio
2

char. Feng et al. (2021) explored the effects of CO2
3 , SO4 , NO3 and Cl on the adsorption of phosphate by using oyster shell-modified sheep manure biochar and lanthanumloaded oyster shell-modified sheep manure biochar, respectively. The results show that the addition of CO2
3 increases the adsorption performance by 2.97%, which was due to 2
the combination of Ca2+ and PO3
4 is inhibited under acidic conditions, and CO3 can promote the combination of Ca2+ and PO3
4 in the oyster shell-modified sheep manure bio

char by increasing the pH of the solution. The addition of SO2
could 4 , NO3 and Cl compete with phosphate ions for adsorption sites and reduced the adsorption capacity

Chapter 7 • Role of biochar in nutrients recovery from wastewater

217

by 9.58%, 8.26%, and 15.05%, respectively. For the lanthanum-loaded oyster shell
2
modified sheep manure biochar, Cl
, CO2
3 , NO3 , and SO4 had no obvious inhibiting effect on the phosphorus adsorption process, and the adsorption amount was only reduced by 0.5%–1%. In addition to coexisting anions, some coexisting cations also have a certain influence on the process of phosphate removal by biochar. Yin et al. (2018c) studied the use of Mg-Al modified soybean straw biochar to recover nitrogen and phosphorus from water. In the presence of NH+4 , the adsorption effect of biochar on PO3
4 was promoted, it may be that Mg2+ on the surface of biochar reacted with NH+4 and PO3
4 to form struvite precipitation. In the study of Li et al. (2017b), it was found that the presence of K+ and Na+ did not affect the adsorption of phosphate by modified sugarcane leaves biochar, while Ca2+ could slightly promote the adsorption of phosphate because it can react with phosphate to produce amorphous calcium phosphate. Other factors In addition to the above factors, the dosage of biochar and the content of modified substances will also affect the removal of phosphorus in water by biochar. In general, within a certain dosage range, the removal rate of phosphate in the water rises with the increase of biochar dosage, because the increase of biochar dosage can increase the total amount of phosphate adsorption sites in the solution (Liao et al., 2018a; Nakarmi et al., 2020). Some research results also show that the excessive dosage of biochar can also have a negative impact on the adsorption effect. For example, in the study of Deng et al. (2021), when the dosage of AlCl3-modified corn cob biochar was increased from 0.03 to 0.15 g, the phosphate removal rate rose sharply, from about 20% to nearly 70%, but when the dosage was further increased, the removal rate showed a downward trend and fell below 30%. This is mainly due to the further increase in the dosage enhancement of the deprotonation of the solution, change the pH of the solution, as well as enhance the electrostatic repulsion between the biochar and the phosphate anion. The content of modified substances on biochar can also promote phosphate removal efficiency to a certain extent. For example, Zhu et al. (2020) found that the removal rate of phosphate by biochar containing nanoMgO particles raised with the increase of the magnesium content on the biochar. When the Mg content on the biochar increased from 0 to 14 wt%, the adsorption capacity increased rapidly, when the content exceeds 14 wt%, the increase rate slows down and finally stabilizes, and the removal rate reaches about 99%.

3.2.3 Removal mechanism Different types of biochar adsorb phosphate in water by different mechanisms, and the adsorption process is often affected by the joint action of many mechanisms. According to the existing research, the mechanism can be divided into the following four categories: electrostatic adsorption, ion exchange, ligand exchange, and surface precipitation (Fig. 3).

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Current Developments in Biotechnology and Bioengineering

FIG. 3 Possible mechanism of phosphate adsorption by biochar.

Electrostatic attraction Electrostatic attraction is one of the important mechanisms of phosphate removal by biochar. The electrostatic effect in the adsorption process mainly depends on the pH of the solution and the zero charge point on the surface of the biochar. When the pH of the solution is less than pHzpc, the surface of the biochar is positively charged and can electrostat2
3
ically adsorb the anionic phosphate (H2PO
4 , HPO4 , PO4 ). However, when the pH of the solution is greater than pHzpc, the surface of biochar is negatively charged, and electrostatic repulsion is generated between the biochar and phosphate anions, thus inhibiting the removal efficiency of biochar to phosphate. In the study of Yang et al. (2018b), when pH < 6, the zeta potential of modified biochar was positive, and electrostatic attraction helped biochar adsorb H2PO
4 in water. With the increase of pH value of the solution, the surface of biochar was gradually negatively charged, and the electrostatic repulsion between phosphate and biochar was enhanced, leading to the decrease of phosphate removal efficiency. Wang et al. (2020) explored the phosphate adsorption effect of a phosphogypsum-modified distillers grains biochar, they found that the biochar got a better adsorption efficiency when the initial pH of the solution was between 3 and 6, and the

Chapter 7 • Role of biochar in nutrients recovery from wastewater

219

phosphate removal rate reached the maximum of 96.7% when the pH ¼ 6 and continued to increase the pH, the removal rate began to decline. This is because the zero-point charge of the biochar used in the study is 7.74, when pH < pHzpc, the surface of the biochar is positively charged, which will generate electrostatic attraction with phosphate ions existing in the form of anions. In addition, metal-containing biochar is readily protonated at low pH (MdOH + H+ ¼ MdOH+2 ), which results in a positive charge on the adsorbent surface, while deprotonation occurs at high pH (MdOHdH+ ¼ MdO
), which results in a negative charge on the adsorbent surface (Xu et al., 2019b). This process can also affect the electrostatic interaction between biochar and phosphate, thus affecting the adsorption effect of biochar on phosphate. Alhujaily et al. (2020) added Mg-Fe modified biochar at pH ¼ 3, and at that pH level, the phosphate removal rate was the highest, up to 79%, after adsorption, the pH rose to 8.6, indicating that MgO and Fe3O4 on the surface of biochar were protons into MgOH+ and FeOH+ through electrostatic interaction, which was beneficial to the adsorption of phosphate. Ion-exchange Ion exchange mainly occurs between the dOH groups on the surface of biochar and phosphate, which is an important mechanism for the adsorption of phosphate by biochar, and mostly occurs in the adsorption process of phosphate by layered double hydroxides (LDH) modified biochar. LDH is a hydrotalcite-type anionic layered compound, which is composed of stacked positive ion layers, the anions in the interlayer channels can be exchanged with various anions (Vithanage et al., 2020). By comparing the XRD and XPS spectra of LDH-biochar composites before and after the adsorption of phosphate, Wan et al. (2017) found that the main reason for the adsorption of phosphate was the intercalation ion exchange with Cl
in the composites. Yang et al. (2019) analyzed the changes of XPS spectra of LDH-biochar composites before and after adsorption and found that the number of hydroxyl groups on the materials decreased after adsorption. The main reason was that hydroxyl groups were replaced by phosphate ions through ion exchange in the process of adsorption of phosphate. In addition to LDH modified biochar, ion exchange is also the mechanism for other types of biochar to adsorb phosphate. For example, in the study of Nakarmi et al. (2020), the hydroxyl on the surface of zinc oxide betaine modified biochar can exchange ions with phosphate, which promotes the removal efficiency of phosphate, the ion exchange process can be expressed as 2Zn-OH + ZnCl2 + 3H2PO4 ! 3Zn-H2PO4 + 2Cl
+ 2OH
. Ligand exchange The ligand exchange mechanism in the process of phosphate removal by biochar refers to that the functional groups (such as dOH and dCOOH) on the surface of biochar form complexes with phosphate ions, or the chemical bonds on the surface of biochar coordinate with phosphate anions to form monodentate and bidentate complexes. This mechanism exists widely in the process of phosphate removal by biochar. In the study of phosphate adsorption by biochar, Liao et al. (2018b) found that the lanthanummodified biochar has a relatively large pH range, and the ionic strength has little effect

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on the adsorption of phosphate by the lanthanum-modified biochar. It is speculated that it may be caused by the formation of the inner sphere complex due to the ligand exchange mechanism, and the subsequent XPS images also confirmed the production of the La-O-P inner-sphere complex. At the same time, the study also found that the pH of the solution after adsorption was higher than the initial pH, which was mainly caused by the release of hydroxide from the hydroxyl group on the biochar during the ligand exchange process. When Ajmal et al. (2020a) studied the interaction between magnetic wood, rice husk biochar, and phosphate, they found that the ligand exchange between the hydroxyl group on the surface of adsorbent material and PO3
4 in liquid can
be expressed as “biochar-Fe(OH)3 + PO3
! biochar-FePO + 3OH .” Yang et al. (2019) 4 4 analyzed the FTIR spectra of LDH-biochar composites before and after adsorption and found that after adsorption of phosphate, a new peak of PdO and P]O appeared at 1050 cm
1, indicating that MdOH groups on the composites formed MdOdP through ligand exchange. Surface precipitation The reaction of metal ions, metal oxides, or hydroxides on biochar with phosphate in water to generate precipitation is considered to be an important mechanism for phosphate removal, which is involved in most of the studies on phosphate removal with biochar. For example, Yao et al. (2013) adsorbed phosphorus in water by MgO-biochar nanocomposite prepared by anaerobic digestion of biomass, and carried out XRD diffraction analysis on the adsorbed biochar material, the results showed that there were Mg-P crystals on the surface of adsorbed biochar material, namely MgPO4 and Mg(H2PO4)2. Therefore, the authors attributed the phosphate removal effect partly to the surface precipitation process between phosphate ions and the MgO and Mg(OH)2 particles carried on the surface of biochar. Li et al. (2020c) used nano-CaO-biochar composite material to remove phosphate in water. By analyzing the XRD and FTIR spectra of the composite material after adsorption, it was found that CaHPO42H2O precipitated on the surface of the composite material, indicating that the precipitation process of Ca2+ and HPO2
4 played an important role in the phosphorus removal process. Xu et al. (2018) used magnesium-modified sawdust biochar to simultaneously recover nitrogen and phosphorus in urine, the main mechanism of this application is that the struvite precipitation is formed through the reaction of Mg2+ on the surface of biochar with NH+ and PO3
4 .

3.3 Sustainable assessment of nutrient recovery The application of biochar to remove and recover nitrogen and phosphorus in water should conform to the concept of sustainable development, which is mainly reflected in the following aspects. First of all, most of the raw materials of biochar come from agricultural and forestry wastes, such as straw, wood, peel, animal manure, etc., and some come from municipal sludge and industrial wastes. These wastes are not only large in quantity, but also difficult to deal with, and often cause serious impacts on the environment due to improper

Chapter 7 • Role of biochar in nutrients recovery from wastewater

221

treatment. For example, most straw biomass disposal methods are incineration, but this treatment method will release a large number of greenhouse gases and generate a large amount of soot, which will cause adverse effects on the atmosphere, soil, and water environment. If these waste biomass are made into biochar, it can not only solve the storage and treatment of such waste, reduce the adverse impact on the environment, but also realize the reuse of waste resources, which is conducive to carbon sequestration and reduction of greenhouse gas emissions. Second, phosphorus is a nonrenewable resource and is widely used. Using biochar made of waste biomass to recover phosphate in wastewater can not only alleviate the eutrophication caused by excessive phosphorus in water, maintain the balance of the aquatic ecosystem, but also facilitate the sustainable utilization of phosphorus and alleviate the shortage of phosphorus resources in the world today. Finally, most of the biochar after adsorbing nitrogen and phosphorus can be reused after desorption, which can reduce the cost of water pollution treatment and improve the efficiency of the material. In addition, nitrogen and phosphorus are indispensable elements for plant growth, so biochar adsorbed with nitrogen and phosphorus can be put into the soil as a slow-release fertilizer, which is not only beneficial to promote seed germination and plant growth and development but also has great benefits for agricultural development. Bai et al. (2017) applied cattail biochar saturated with ammonia nitrogen to soil conditioning, and showed good water retention, nutrient retention, and supply capability. Chandra et al. (2020) used column leaching experiments to explore the effects of KOH and FeCl3 modified rice straw biochar on the transport and leaching of soil nutri+ 3
ents (NO
3 , NH4 , and PO4 ) and found that biochar not only keeps nutrients attached to the soil surface but also it can release nutrients in phases and in time to improve the utilization rate of nutrients by plants. Li et al. (2020c) used nano-CaO2 modified wheat straw biochar to adsorb phosphate and put it into the soil for tomato seedling growth experiment, the results showed that phosphorus-containing biochar could promote the growth of tomato seedlings to some extent. The application of biochar in soil improvement has achieved good effects, but many application values of biochar have not been explored. Therefore, it is necessary to develop a variety of application approaches of biochar and develop biochar with different efficiency according to different application values.

4. Conclusion and perspectives This chapter introduces the preparation methods and physicochemical properties of different biochar. The properties of biochar prepared from different raw materials are quite different. Pyrolysis temperatures also significantly affect the surface area and functional groups of the biochar. Modifications with acids, oxidants, metal salt/oxide, and other composite materials usually can increase the surface area and oxygen-containing functional groups of biochar, resulting in the enhanced nitrogen and phosphorus removal and recovery capacity. In this chapter, the influence factors and removal mechanism of

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Current Developments in Biotechnology and Bioengineering

biochar on removing nitrogen and phosphorus from water are discussed and analyzed as well. The preparation method of biochar, pH of the aqueous solution, and coexisting ions all dramatically affect the adsorption of nitrogen and phosphorus by biochar. The multiple mechanisms are usually involved in the removal of nitrogen and phosphorus by biochar in water. The removal mechanisms of nitrogen are mainly electrostatic attraction, ion exchange, and surface complexation, while the removal mechanisms of phosphorus are mainly electrostatic attraction, ion exchange, ligand exchange, and surface precipitation. As an effective and environment-friendly adsorbent, biochar shows great potential in removing and recovering nitrogen and phosphorus in the water. However, the scalingup application of this novel and environment-friendly material in the real practices faces several challenges. (1) Although the modification can significantly improve the adsorption performance of biochars, the preparation and modification process of biochar is relatively complicated and most modification processes also produce secondary pollution. Therefore, the economical and environment-friendly biochar processing and modification is needed. (2) Considering the significant difference in the physicochemical properties and adsorption performance for biochars prepared from different raw materials and preparation method, it is necessary to establish unified raw material standards, biochar product indicators, performance evaluation standards, and methods to expand the popularization and application of biochar. (3) The current research mainly focuses on the adsorption performance of biochar, the regeneration and reuse of biochar should be explored to give full play to the resource value of biochar, such as adding desorption and cyclic adsorption experiments or designing a continuous flow reactor in the researches. (4) The research on the removal of nitrogen and phosphorus from water by biochar is conducted in the laboratory. However, the natural water or wastewater is complex, and the coexisting substances will have different impacts on biochar, and some even lead to the desorption of the pollutants. Therefore, it is necessary to explore the removal and recovery performance of biochar for nitrogen and phosphorus in actual water bodies, especially developing highly selective adsorption methods. (5) There are many modification methods of biochar, and the removal mechanism of pollutants by different modified biochar is also different. The removal mechanism of biochar prepared by different modification methods also needs to be further explored. (6) It is worth noting that biochar may have adverse effects on the environment, the potential environmental risk of biochar application should be evaluated.

References Abdul, G., Zhu, X., Chen, B., 2017. Structural characteristics of biochar-graphene nanosheet composites and their adsorption performance for phthalic acid esters. Chem. Eng. J. 319, 9–20. Ajmal, Z., Muhmood, A., Dong, R., Wu, S., 2020a. Probing the efficiency of magnetically modified biomassderived biochar for effective phosphate removal. J. Environ. Manag. 253, 109730. Ajmal, Z., Muhmood, A., Dong, R., Wu, S., 2020b. Probing the efficiency of magnetically modified biomassderived biochar for effective phosphate removal. J. Environ. Manag. 253, 109730.

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8 Application of biochar for improving sewage sludge treatment Bing-Jie Ni, Lan Wu, Xingdong Shi, and Wei Wei CENTRE FOR TECHNOLOGY IN WATER AND WASTEWATER, SCHOO L O F CIVIL AND E NVIRONMENT AL ENGINEERING, UNIVERSITY O F TECHNOLO GY SYDNEY, SYDNEY, NS W, AUSTRALIA

1. Introduction Huge amounts of waste sludge are produced from wastewater treatment plants (WWTPs) daily, imposing a striking challenge of valorizing this organic waste as latent resources (Guo et al., 2020a; Liu, 2019). Additionally, considering conventional energy, a vital part of a daily life, will be diminished soon or later, finding out a cost-effective and environment-friendly technology to optimize the utilization of sewage sludge as energy carriers has never been more important than today (Qiu et al., 2019a). Anaerobic digestion (AD) is a well-documented technology to achieve the goal of “Turning Waste into Wealth.” In spite of great efforts have been dedicated to achieve this target using AD as the principle approach, the high moisture content in sewage sludge is a key issue that is needed to be addressed to lower down the cost of sludge treatment and transport (Zhen et al., 2018). Nevertheless, due to the limitation imposed by extracellular polymers substances (EPS) in sludge, the interstitial water containing in this organic waste cannot be removed sufficiently. High moisture content of the sludge will then hinder the application of this organic waste as energy resources. Luckily, the addition of biochar, a charcoal created by the pyrolysis of biomass, has been found to facilitate the dewatering of sludge. In detail, the existed persistent free radicals possessed by such coal-like material are effective oxidizer to destroy sludge floc and the EPS of organisms, releasing interstitial water thereafter (Abelleira et al., 2012). Thus, to further explore the potential of biochar in enhancing the performance of dewatering sludge, a comprehensively evaluation concerning the sustainable application of biochar in improving sludge dewaterability is needed. Great efforts have been carried out to convert the organic matter in sludge into either energy resources (i.e., H2 or CH4) or valuable resources such as volatile fatty acids (VFAs) through AD (Shen et al., 2015; Yu et al., 2019). Biological H2 production process was deemed to be more feasible than chemical-physical methods, as which can operate under simple operational conditions to yield steady H2 yield with only a little energy consumption (De Gioannis et al., 2013). Strategies have been proposed to promote this fossil fuel alternative, including adding biochar. High amount of CH4 containing in biogas is another Current Developments in Biotechnology and Bioengineering. https://doi.org/10.1016/B978-0-323-91873-2.00007-8 Copyright © 2023 Elsevier Inc. All rights reserved.

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typical energy resource obtained from sludge digestion. However, the CH4 content in biogas yielded from single phase AD is less than desirable due to high CO2 content, inducing low specific energy value and, therefore, poor ignition quality subsequently (Sunyoto et al., 2016). Hence, researchers have dosed waste-driven biochar into the digester to improve associated methane content due to the specific features of carbon-rich material such as large specific surface area, high porosity, small bulk density, high stability, feeble electrical conductivity (Chiappero et al., 2020). The yield of VFAs, prerequisite of CH4, can also be enhanced by supplying biochar while digesting sludge. Given such carboxylates can also act as food addictive and industrial raw materials (Hanczakowska, 2017; Xu et al., 2015), it is necessary to strike for higher VFAs generation while producing H2 and CH4. Collectively, adding waste-driven biochar is a feasible way to convert sewage sludge into wealth by promoting higher H2, CH4, and VFAs production. A relatively comprehensive evaluation on the impacts of biochar addition on sewage treatment is important to intensify the waste management via AD. Therefore, this review will handle this issue from the aspects of (1) evaluating the dewatering performance of sludge after dosing biochar in terms of associated EPS alternation and latent mechanisms; (2) discussing the role of biochar in enhancing CH4 while digesting sludge; (3) understanding the mechanisms of applying biochar for producing higher H2 and VFAs from AD; and (4) introducing the effects of biochar on sludge reduction. The outlooks concerning the function of biochar in achieving the sustainable sludge treatment are then finally proposed.

2. Sustainable application of biochar in improving sludge dewaterability High amount of waste sludge is produced from WWTPs due to the continuous growth of microorganisms. The important feature of this waste sludge is huge volume resulted from its extremely high moisture content (typical over 92% moisture by weight), increasing the costs of sludge treatment thereafter (Anjum et al., 2016; Nellenschulte and Kayser, 1997). Reducing sludge volume through dewatering process is an attractive method to lower the operation costs of the subsequent treatment. In general, the categories of water containing in sludge can be divided into four parts, free water, interstitial water, surface (or vic
zac, 2004). Mechanical inal) water, and bound (or hydration) water (Vaxelaire and Ce dewatering (e.g., filtration and centrifugation) can easily remove free water and a part of bound water. The remains of bound water inside the cell structure can be further released via disrupting sludge cells—called conditioning process (Erdincler and Vesilind, 2003). Recently, advanced oxidation processes (AOPs) including Fenton and Fenton-like processes have been recognized as the most efficient chemical conditioning to release bound water via destroying sludge flocs by free radicals (He et al., 2016; Neyens et al., 2004). It has been proven that biochar can catalyze the decomposition of oxidants into free radicals to enhance the removal of pollutants. Therefore, biochar has also the same role in stimulating the release of free radicals leading to enhanced dewatering.

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The objective of this section is to describe the principal and performance of the biochar in sludge dewatering. To understand the role of biochar in sludge dewatering, the first two sections will introduce the interaction mechanisms and the EPS variation. The dewaterability performance of biochar will be presented lastly.

2.1 Interaction mechanisms 2.1.1 Radical generation In general, free radicals generating from oxidants (hydrogen peroxide or persulfate) can destroy sludge flocs to stimulate releasing bound water (He et al., 2016; Neyens et al., 2004). Commonly used strategies for enhancing radical generation include chemical activation [alkaline (Qi et al., 2016) and benzoquinone (Zhou et al., 2015)], transition metal activation [e.g., Fe, Co, Ni, etc. (Anipsitakis and Dionysiou, 2004)] and energy activation (heat, ultrasonic and ultraviolet) (Zhao et al., 2021b). Recently, biochar and biochar-based materials are extensively investigated in improving radical generation (Huang et al., 2020; Ouyang et al., 2019; Xu et al., 2020). Active nanoparticles can be loaded on the biochar to synthesize the modified catalyst (Xu et al., 2020). In this case, these active nanoparticles are the main activators. In fact, except for being a catalyst carrier, biochar can also catalyze % persulfate or hydrogen peroxide to produce sulfate radicals (SO
4 ) or hydroxyl radicals % ( OH) (Pan et al., 2021). In this section, two roles of biochar, catalyst and support will be described in detail. Catalyst: The mechanism of biochar to active hydrogen peroxide is the higher generation of %OH caused by electron transference (Pan et al., 2021). The activation ability of biochar is influenced by transferred electron density and available cross-linked microporous structure (Fang et al., 2014). Heterogeneous metal atoms and free radicals (PFRs) groups (e.g., semiquinones, and phenoxyls) contained in biochar are responsible for catalytic ability of biochar (Devi et al., 2016). Studies suggested that the formation of PFRs is attributed to the thermal decomposition of phenols, hydroquinones and other organic compounds in the presence of metal oxides (Fang et al., 2015a). PFRs are highly reactive to catalyze the decomposition of hydrogen peroxide to form %OH. The process of generating free radicals by PFRs in hydrogen peroxide system can be summarized as the follows. The oxygen in the water obtains electrons from biochar to form superoxide radicals (%O
2 ), which then accepts an electron from PFRs to generate hydrogen peroxide. The resulting hydrogen peroxide would then react with PFRs to produce %OH (Fang et al., 2015b). Similar to hydrogen peroxide system, the free radicals driven from biochar-catalyzed persulfate conversion are also due to electrons transfer (Eqs. 1, 2). The formation of free radicals is attributed to the following activation sites (Fig. 1), (1) O-functional groups on biochar surface, including carbonyl (dC]Od), hydroxyl (dOH), and carbonyl (dCOOH) (Kemmou et al., 2018); (2) the PFRs on biochar surface (Fang et al., 2015a); (3) the internal and edge sites on the defective structure of biochar. Such special structure is formed during pyrolysis and usually exhibited as graphite or unsaturated carbon (Ouyang et al., 2019); (4) sites on biochar where π-π transition occurs (He et al., 2019);

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FIG. 1 The mechanisms of biochar catalyzing persulfate to form free radicals. Originally designed by Pan et al. (2021).

(5) the electron holes formed when biochar provides electrons (Duan et al., 2018); and (6) ammonia or organic nitrogen and metal ions sourced from the sludge-derived biochar (Yin et al., 2019).
 2

Persulfate ðPSÞ activation : S2 O2
8 + e ! SO4 + SO4

(1)




Peroxymonosulfate ðPMSÞ activation : HSO
5 + e ! SO4 + OH

(2)

Support: Biochar can be also applied as the catalyst support. Fig. 2 shows the main roles of biochar as support to generate free radicals. It can stabilize and disperse the nanoparticles to prevent metal from releasing to the aqueous phase (Faheem et al., 2020). In addition, the number of active sites is also increased due to the expanding surface area (Chen et al., 2018). Studies indicated that the surface area of the prepared biochar composite with nanoparticle addition is larger than that of the original nanoparticles along (Manikmian and Liu, 2018). More importantly, electron transfer or shuttling electrons would be enhanced by regulating electron-transfer reactions among PMS and PDS via biochar, as such coal-like material supplies the electrons to the reaction due to its own catalytic ability (Wang et al., 2019).

2.1.2 Dewatering The radicals generating from biochar/H2O2 or biochar/persulfate can oxidize sludge floc and extracellular polymeric substance (EPS). The presence of EPS can lead to the difficulty in sludge dewatering due to the significant role of EPS playing in binding water (Neyens and Baeyens, 2003). After destroying the sludge floc and EPS, the interstitial water existing in the sludge flocs and water (bound water) combined with EPS are released from sludge

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233

FIG. 2 The schematic diagram of biochar support. Originally designed by Zhao et al. (2021b).

(Abelleira et al., 2012). EPS has high affinity for water, resulting in the increment of sludge viscosity when EPS content increases. Thereby, the sludge containing high level of EPS is hard to dewater (Zhen et al., 2012a). When EPS contacts with radicals (%OH or SO
 4 ), flocs are destroyed immediately. The interstitial and vicinal water containing in the space of flocs will be then transformed into free water, improving the sludge dewaterability greatly. EPS variations are responsible for this enhanced sludge dewatering, which will be described in Section 2.2 in detail. Notably, the penetration of radicals into cell walls would lead to increased permeability of the cell membranes, thereby improving the release of intracellular water (Zhen et al., 2012b). In short, the higher generation of radicals resulted from biochar addition promotes the release of intracellular water and bound water, improving sludge dewaterability consequently. In addition, except for enhancing dewaterability via EPS degradation by radicals, biochar can also stimulate sludge cake to form porous structure (Guo et al., 2019; Guo et al., 2020c; Wu et al., 2016). This porous structure is a kind of permeable and rigid lattice structure that will ensure an unblocked water filtration channel even under high pressure (Qi et al., 2011). Current dewatering processes are commonly mechanical processes including centrifugation, vacuum filtration and pressure filtration. The dewatering rate of these processes is often constrained by highly compressible sludge blocking the water canal (Qi et al., 2011). In this case, longer compression time or higher pressures are needed to further improve the dewatering efficiency. Therefore, biochar is used to reduce the compressibility of sludge and improve the mechanical strength along with the permeability of the sludge solid during compression. The role of biochar is usually named as skeleton builders or filter aids. However, skeleton builders cannot destroy the structure

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of EPS along. Normally, these skeleton builders are always combined with AOPs which could generate radicals to degrade EPS for releasing bound water.

2.2 EPS variations EPS are regarded as the most important factor in limiting sludge dewatering (Neyens and Baeyens, 2003). They are mostly composed of polysaccharides, proteins and other natural polymers with high molecular weight. EPS surrounding the bacterial cell wall is thought to be a protective gel-like reservoir for water preservation (Abelleira et al., 2012). Previous studies indicated that the concentration of EPS would increase zeta-potentials with low contact angles (high hydrophilicity) (Pere et al., 1993), which prevents efficient flocculation. Biochar treatment based on oxidation of radicals influences the EPS components, leading to the enhanced sludge dewaterability thereafter. Therefore, comprehending the EPS variations after biochar/oxidants treatment is essential to understand the enhanced dehydration performance.

2.2.1 Total organic carbon Some studies found that the sludge dewatering performance is influenced by the contents of hydrophobic organic substances in EPS (Guo et al., 2020b; Ren et al., 2015). In general, the higher proportion of hydrophobic organics in EPS means greater difficulty in sludge dewatering. Some researchers indicated that the content of total organic carbon (TOC) in EPS decreased after treating sludge by biochar/PS system (Ren et al., 2015). This decrease is mainly manifested as the reduction of hydrophobic acidic substance (HPO-A), hydrophobic neutral substance (HPO-N) and transitional hydrophilic acidic substance (TPI-A) fractions of EPS (Ren et al., 2015). A higher decrement will occur when PS concentration increases. In contrast, transitional hydrophilic neutrality matter (TPI-N) and hydrophilic substance (HPI) fractions of EPS has no obvious change (Ren et al., 2015). Overall, the enhanced sludge dewaterability in the biochar/oxidants system is attributed to the decrease of hydrophobic organics.

2.2.2 Protein and polysaccharide It has been confirmed that the protein and polysaccharide component in EPS is the principal cause for the difficulty in sludge dewatering (Liu et al., 2016; Zhen et al., 2012a). After biochar/PS oxidation treatment, the total content of protein and polysaccharide in EPS are reduced to varying degrees. The content of protein in soluble (S-), loosely bound (LB-) and tightly bound (TB-) EPS is disrupted efficiently (Guo et al., 2020b). Guo et al. (2020b) found that the trend of polysaccharides in LB- and TB-EPS was similar to the alternation of protein after supplying biochar, but the amount of polysaccharide in S-EPS increased due to the disruption of bound EPS to S-EPS. Other studies pointed out that TB-EPS were main reason causing the difficulty for sludge dewatering (Yuan et al., 2014). Therefore, it is reasonable to conclude that the degradation of TB-EPS by biochar/oxidation system can enhance the release of EPS-bound water and dewaterability of sludge.

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2.3 Dewaterability performance The main indexes to evaluate the sludge dewaterability include capillary suction time (CST), specific resistance to filterability (SRF) and water content (WC). Guo et al. (2020b) applied corn biochar to active PMS generating radicals to enhance waste activated sludge dewatering. After treatment, the moisture content (MC) decreased from 95.5% to 43.4% with 4.21 times increase in standardized-capillary suction time (SCST) (CST reduction of 76.80%). Another study also observed enhanced sludge dewatering by iron-rich biochar-driven Fenton reactions (Tao et al., 2019). In this research, the reduction percentage of the CST and SRF is 90.24% and 98.01%, respectively at the doses of 0.792 g/g vs. biochar and 0.072 g/g vs. H2O2. Moreover, this process can reduce 28.39% total operating cost of sludge treatment compared with other methods (Tao et al., 2019). Some studies also evaluated the sludge dewatering efficiency of modified biochar as the skeleton builders. Although skeleton builders can form porous structure in sludge cake to promote sludge dewaterability, it cannot destroy the structure of EPS which is the main difficulty of dewatering. Therefore, current studies often modified biochar or compared biochar with AOPs. For example, sludge-based biochar could enhance skeletal support and improved the net sludge solids yield by 138% (Guo et al., 2020c). Wu et al. (2016) also used rice husk biochar modified by FeCl3 to promote sludge dewatering. The results showed that SRF and WC were decreased by 97.9% and 18% respectively compared with the control group (Wu et al., 2016).

3. Sustainable application of biochar in improving anaerobic sludge digestion Energy and valuable resources can be recovered from solid wastes via AD cost-effectively and environmental-friendly (Yu et al., 2019). Methane (CH4) is the main product of AD, accounting for 50%–70% of total biogases (Shen et al., 2016). Supplementation of biochar is thought to be a promising and efficient way to enhance process performance and maintain the stability of AD (Shen et al., 2017). Biochar plays various roles in AD to encourage the recovery of carbon of wastes in the form of biogas or value-added products. In detail, benefiting from its porous structure, biochar could reduce the inhibitory effect of ammo€ et al., 2016; Luo et al., 2015; nia and provides a place for methanogens attachment (Lu Mumme et al., 2014). Additionally, an “electron shuttle” could be formed to promote interspecies electron transfer (IET) between methanogens and other microbial communities by biochar, named as direct interspecies electron transfer (DIET). To fully understand the mechanism and performance of enhanced AD by biochar, the enhancement of methane production and how biochar serves as “electron shuttle” to establish DIET will be described in the first two sections. The variations of the microbial community will be presented in the last section.

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3.1 Enhancement of methane production 3.1.1 Performance AD process can be divided into three steps, hydrolysis, acidification, and methanogenesis (Inthapanya et al., 2012). Methanogenesis is the most sensitive stage among all AD processes. A small deviation from operational parameters would negatively impact methane production (Li et al., 2018). Previous studies indicated that biochar could significantly promote AD reactor to produce CH4. Table 1 exhibited the improvement of CH4 yield by adding biochar to AD. The first research concerning the introduction of biochar into AD to improve biogas yields was conducted by Inthapanya et al. (2012). The biogas production rate in this study was increased by 31% in their study (Inthapanya et al., 2012). After then, various studies focused on improving the biogas or methane production rate in AD reactor. Luo et al. (2015) used biochar based on fruitwoods to promote AD in an attempt to increase CH4 yield up to 86.6%. Except for increasing CH4 yield, biochar could also improve the methane volume percentage in the biogas of AD. For example, the biochar-amended digesters could obtain 92.3% average methane content in biogas, while this percentage is only around 77% in the control group (Shen et al., 2016). According to the report, biochar could also improve the stability of methane production in AD process.

Table 1

The efficiency of biogas/methane yield by the addition of biochar.

Biochar feedstock

Production method

Substrate

Temp.

Biogas/CH4 yield

Ref.

Rice husks

Pyrolysis

Cattle manure

35°C

Biogas yield increased 31%

Rice straw Orange peel Bamboo powder Fruitwoods

Pyrolysis

Rice paddy soil

30°C

CH4 yield decreased 20.8% CH4 yield decreased 17.0%

(Inthapanya et al., 2012) (Fu et al., 2021)

Pinewood

CH4 yield decreased 15.8% Pyrolysis at 800°C Gasification

Pulp sewage/glucose

35°C

CH4 yield increased 86.6%

Water reclamation plant sludge

37°C

Purchased

Mixture of primary sludge and WAS

55°C

CH4 yield increased 2.9% (CH4 content of 92.3% in biogas) CH4 yield increased 7.9% (CH4 content of 89.8% in biogas) CH4 yield increased 4.3% (CH4 content of 79.0% in biogas) CH4 yield increased 4.0% (CH4 content of 78.5% in biogas) CH4 yield increased 36.9%

Pyrolysis at 400°C

Aqueous pyrolysis liquid

40°C

CH4 yield increased 16.6% CH4 yield increased 60%

White oak Pinewood

55°C

White oak Corn stover Pine Corn stalk pellet

(Luo et al., 2015) (Shen et al., 2016)

(Shen et al., 2017) (Torri and Fabbri, 2014)

Chapter 8 • Application of biochar for improving sewage sludge treatment

237

After applying charcoal, the stability of reactor could be maintained at high COD loading rate (2.17 g COD/L/d), and the methane production rate was 1.6 times higher compared to the control without charcoal (Watanabe et al., 2013b). Biochar seems to have a pronounced improvement in mesophilic digesters compared with thermophilic digesters (Shen et al., 2016). This phenomenon is attributed to the weak adaptability of microorganisms in the initial inoculum to the temperature changes (Bousˇkova´ et al., 2005).

3.1.2 The role of biochar in methane production The features of biochar including large specific surface area, rich porosity, and high electrical conductivity allows such charcoal to obtain the ability in removing CO2, H2S, and NH3, relieving the negative effects imposed by those gases on methanogens subsequently (Zhang and Lu, 2016). For example, the HCO3
/CO32
transformed from CO2 could buffer the pH conditions by reacting with alkali and alkaline compounds released by biochar (Shen et al., 2016). In addition, excess ammonia accumulation would inhibit AD of nitrogen-rich wastes such as animal wastes and slaughterhouse by-products. More crucially, biochar could establish DIET to improve the efficiency of electron transfer (Wang et al., 2018c). DIET will be discussed in detail from the perspective of its establishment and mechanism in Section 3.2. Overall, biochar plays various roles in CH4 production by working as inhibitor for adsorption, pH buffer for adjusting alkalinity in the reactors, promoter for improving electron transfer efficiency and microbial enrichment selector for optimizing the structure of reactor microbe. Following context will strengthen the role of biochar in CH4 production with regards of IET, microbial community and the specific function of biochar in recovering energy from AD. During the AD process, some VFAs could be produced as intermediates by acidforming bacteria and reduce pH level. Methanogens convert these carboxylates to methane, balancing the pH level in the AD reactor thereafter (Zhang et al., 2014). However, VFAs would accumulate once high organic loads are fed into the system, because the VFAs production rate exceeds the consumption rate under such cases. The pH drop would then be observed and inhibit the biosynthesis of CH4. It has been verified that biochar is an efficient material to improve the buffer capacity of the system. For example, the impacts of biochar on AD process were widely investigated. The results showed that the biochar increased the total alkalinity from 4800 to 6800 mg/L CaCO3, enhancing the buffer capacity of the AD system. Another research also observed enhanced methane production with a lower total VFAs concentration (Jang et al., 2018). This performance is mainly due to its functional groups and inorganic components (Chiappero et al., 2020). Functional groups (e.g., amine) above biochar react with proton, thereby mitigating the pH reducing, while the inorganic portion like alkali and alkaline earth metals are also responsible for consuming proton (Jang et al., 2018). Another important role of biochar in promoting methane production is that it can serve as an adsorbent to remove harmful chemicals. Several chemicals are responsible for inhibiting AD process and reducing bio-methane yields, such as metals (Cu2+, Zn2+, Cr3+, etc.), organic matter (e.g., pesticides and antibiotics), sulfides and ammonia.

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Ammonia and ammonium have attracted more attention compared to other chemicals € et al., 2016; Su et al., 2019), as excess free ammonia nitrogen con(Ciccoli et al., 2018; Lu centrations is the major cause for AD failure, especially when nitrogen-rich substrate is the only substrate for digestion (Rajagopal et al., 2013). Biochar addition can enhance the tolerance of AD to the high concentration of ammonia. For example, the addition of biochar could alleviate inhibition of ammonia under 1.5 g-N/L ammonia conditions when fermenting food wastes anaerobically. Giwa et al. (2019) reported that biochar addition can alleviate the inhibition caused by high ammonium-N concentration (>2450 mg/L), while benefiting the biodegradation of VFAs and promoting the breeding of archaea in longrun operations (Giwa et al., 2019). However, the mechanisms for free ammonia mitigation are still unclear with ongoing investigation required (Chiappero et al., 2020). The main hypotheses for explaining the alleviation of negative impacts caused by biochar addition up to now include adsorption, cation exchange, and reaction of surface functional groups (Chiappero et al., 2020). Literature suggested that the biochar exhibits promising values of ammonium adsorption capacity. For example, Qiu et al. (2019b) pointed out that the rage of ammonia sorption is 8.61–114 mg N/g. Although large porous structure with high specific surface area could result to this high sorption ability, some studies still found other adsorption mechanism including ion exchange between ammonium and acidic functional groups above biochar (Kizito et al., 2015; Takaya et al., 2016). Therefore, alleviating ammonia inhibition by biochar may be a complex physical and chemical process. More efforts are still needed to reveal the interactions between biochar and ammonia.

3.2 Establishment of direct interspecies electron transfer Plenty of studies indicated that the addition of biochar could improve electron transfer efficiency among syntrophic partners. This electron transfer among various microorganisms is called IET. In AD process, acetogens and methanogens commonly established a syntrophic connection for achieving a stable CH4 production. The acetate is utilized by acetate-consuming methanogens to produce H2, while the hydrogen finally is transformed to CH4 by H2-consuming methanogens (Karakashev et al., 2005). In this methanogenesis, H2 acts as electron carrier between the acetogens and methanogens. This method of electron transfer via H2 or formate is named as indirect interspecies electron transfer. H2/formate have been deemed as the main matter to transfer electron between the acetogens and methanogens for decades (Rasapoor et al., 2020). However, recent studies found that some electrically conductive material such as biochar could directly transfer the electron from one cell to another without H2 and formate, i.e., DIET (Park et al., 2018). It is verified that DIET is more efficient than indirect interspecies electron transfer, as which require less energy consuming and reaction time than last one (Cheng and Call, 2016). Three proposed mechanisms for processing DIET have been published till now (Rasapoor et al., 2020), including (1) conductive materials; (2) electron conduit; and (3) replaced membrane-bound electron transport proteins. No direct evidence has been provided to support the hypothesis regarding the mechanisms for conducting DIET.

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As biochar is one of the conductive materials attracting increasing attention due to its superior physicochemical property and the easy for manufacture recently, many studies have attributed the improvements in methane production to DIET when this coal-like material is presented. However, the functions of biochar to improve IET efficiency via DIET are still in debates. More specifically, although biochar was deemed to promote IET via either chemical-based electron exchange mechanisms or physical-based conductive mechanisms (Wang et al., 2020), the knowledge regarding electron transfer pathway among syntrophic partners with biochar involved are still unclear. Several methods have been performed to confirm the occurrence of DIET in anaerobic systems with biochar involved, such as 16S rRNA gene sequencing, metagenomics, and cyclic voltammetry approaches (Van Steendam et al., 2019). Current researches shown enhanced IET by biochar are mainly based on the detection of bacterial and archaeal species with DIET function using 16S rRNA gene sequencing or metagenomics (Chiappero et al., 2020). For example, DIET between the Methanosarcina barkeri and Geobacter metallireducens was revealed by Zhao et al. (2015). In their research, 86% electrons retrieval as CH4 production from ethanol by biochar addition was observed. In addition, other species such as Clostridium, Geobacter, Coprothermobacter, Thauera, Corynebacterium, Spirochaeta, and Syntrophomonas species could also donate electrons to conductive material (Lee et al., 2016; Lei et al., 2016). Methanogens then accept these electrons from the conductive material resulting in an enhanced CH4 production (Park et al., 2018). However, this method still obtains its own limitation, as current understanding on the organisms involved in DIET during digestion is not comprehensive. Therefore, the microbial consortia involved in DIET process cannot be fully revealed.

3.3 Evolution of microbial community Generally, the most predominant microorganism in anaerobic digester is mainly bacteria instead of archaeal, with the main phyla belonging to Firmicutes, Bacteroidetes, Proteobacteria, Actinobacteria, and Thermotogae (Shen et al., 2017). Although various bacteria were detected in the biochar-medicated AD reactor, the microbial community varies with dif€ et al., 2016; Su et al., 2019; Wang et al., 2018c). In addition, Methaferent researches (Lu nosaeta, Methanobacterium, Methanosarcina, and Methanolinea are the most common archaea species detected in the reactor (Chiappero et al., 2020). As mentioned above, the addition of biochar has positive effects on the microbial community, especially on the DIET-related species. One of the impacts is that biochar could stimulate the proliferation of anaerobic microorganisms in anaerobic digester (Kumar et al., 1987). Biochar also optimized microbial community structure with enriching Methanosarcina spp. (74.9%), while this percentage is only 17% in the nonbiochar reactor (Qiu et al., 2019b). In addition, the number of SRB was also significantly decreased. Inactive SRB means low conversion of SO2
4 to H2S which is harmful for Methanogens. Luo et al. (2015) found that abundant Methanosaeta spp. in the anaerobic digester and biochar can improve the resistance of acidic conditions. Moreover, the microorganisms attached on the large surface area of

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biochar can result in more enhanced random collision among microbiomes, thereby stimulating releasing extracellular polysaccharides and assisting biofilm formation (Sun et al., 2020). Moreover, these tightly clustered microorganisms in biofilms could promote electron transformation and defend against toxic compounds (Liu et al., 2017). Summers et al. (2010) firstly point out the idea of DIET in a coculture of G. metallireducens and Geobacter sulfurreducens (Summers et al., 2010). Other potential DIET functional microorganisms were then found out after Summers’ study. In detail, exoelectrogenic microorganisms could release electrons which were subsequently transferred to methanogens via IET without using H2 or formate (Barua and Dhar, 2017). It is proved that Methanosarcina spp. and Methanosaeta spp. are the main archaea species that can produce CH4 via IET (Stams and Plugge, 2009). Reported syntrophic partners include Geobacter spp. (Stams and Plugge, 2009), Syntrophomonadaceae (Li et al., 2015), Coprothermobacter spp. (Yamada et al., 2015), Thauera spp. ( Jing et al., 2017), Bacillaceae spp. (Zhuang et al., 2015), etc. DIET observed in many studies are usually based on indirect observations through detecting the enrichment of bacterial and archaeal species with DIET function. For instance, higher abundant Geobacter and Methanosaeta were detected in the butyrate and propionate-fed reactor with biochar supplement than the blank, indicating the possible occurrence of DIET in this system (Zhao et al., 2016). Notably, Zhao also detected organism with capability to conduct IET by conducting interspecies H2 transfer in their react, such as Syntrophomonas and Smithella. Another study analyzing the microbial community found the enrichment of potential syntrophic DIET-partners (Wang et al., 2018a). In this research, the related-DIET bacteria and archaeal included Geobacter, Bacteroidetes, Methanosaeta, and Methanosarcina. Strict acetotrophic Methanosaeta is different from Methanosarcina which can use acetate, methanol and other substrates to produce methane (Guo et al., 2014). In addition, Methanosarcina can even operate acetoclastic methanogenesis at a high free ammonia level of 916 mg/L (Hao et al., 2015). Overall, how biochar influences microbial community is still unclear. The potential mechanisms are proposed as follows. Firstly, biochar could supply the environment for microbial attachment and acclimation, accelerating biofilm formation and improving the tolerance of microorganisms for toxic compounds consequently. Secondly, some functional microbes, i.e., DIET-related species, are selectively enriched, indirectly verifying the construction of DIET by introducing biochar. However, this hypothesis should be verified by providing more direct evidence rather than simply supplying microbial sequence detection in future work.

4. Sustainable application of biochar in improving anaerobic sludge fermentation Sewage sludge is rich in organic matter and other easily degradable nutrients for plant growth. However, the direct use of sludge as fertilizer might be unwise, as such kind of

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241

waste may contain high level of toxic heavy metals which can pose threat to both environmental conditions and public health. Therefore, AD has been adopted for years to stabilize sewage sludge by reducing sludge volume and recovering energy in the form of biogas or hydrogen. Nevertheless, this biological process is constrained by the step of yielding VFA, the key prerequisite for CH4 production. Too high or too low VFA production rates will lead to the failure of digestion process by exerting inhibition on methanogens or break the balance among VFAs consumption and CH4 production. The dosage of biochar can address such dilemma to some extent as introduced in Section 4.1. As for hydrogen, such clean energy carrier has been employed for various applications as alternatives for fossil fuels. Compared to other H2 production routes, converting sludge into H2 during AD process is relative more sustainable to yield this renewable energy. Lately, biochar derived from organic wastes has been applied to enhance the biological production of H2. Therefore, in order to enhance the energy recovery of sludge, this section will review the mechanisms and the enhancement of VFA and H2 by anaerobic organisms as below.

4.1 Mechanisms and enhancement of volatile fatty acids production Given VFAs are not only the main precursor for CH4 formation (Qiu et al., 2019a; Zhai et al., 2020) but also the high value-added products (Cavalcante et al., 2017), the said carboxylates are valuable resources recovered from AD. Many researchers have studied the effect of biochar addition on VFAs accumulation from digesters with different or even opposite conclusions been reported (Table 2). This is probably due to the improvement of VFAs yields are closely related to the characteristics of biochar varying with the biomass sources (Lu et al., 2020b). Additionally, pyrolysis temperature adopted in carbonizing biomass can also directly regulate biochar properties including sorption potential and ion-exchange capacity. For instance, increasing pyrolysis temperature would induce larger surface area, as the volatile content in the biochar were largely removed. The impact of charcoal on microbial community of AD also depends on the specific type of biochar. Hence, choosing the right biochar type to digesters is essential to promote the performance of AD by enhancing VFAs production (Zhang et al., 2019). To further extend the application of biochar in carbon recovery from the wastes via AD, understanding the mechanisms and enhancement of VFAs production is essential. The mechanisms of biochar influencing VFAs accumulation are probably due to four main reasons as introduced below. The first explanation for higher VFAs production is the alternation in microbial community structure and related enzymatic activities (Lu et al., 2020a). For instance, the abundance of acid-forming bacteria such as Clostridiales and the organisms working on hydrolysis including Bacillaceae were all reported to increase after adding carbon-based material (Awasthi et al., 2018; Zhai et al., 2020). Higher enzymatic activities were also obtained in the systems amended with algae-driven biochar (Duan et al., 2019). The transformation of organic matter into low-molecular organics was then promoted along with higher VFAs production thereafter (Duan et al., 2019). The porous surface of biochar might be the second reason for VFAs accumulation, as the pores within

242

Current Developments in Biotechnology and Bioengineering

Table 2

Feedstock

VFA production via anaerobic fermentation with biochar addition. Dominant functional groups

Specific surface (m2/g)

VFA titers without biochar addition

VFA titers with biochar addition

Biochar dosage

326.2 mg/L

1017.5 mg/L

10%

(Zhai et al., 2020)

–

40 mg/L

– 0.2–1.5 g/L

(Lu et al., 2020a) (Duan et al., 2019)

0.6 g/L

(Lu et al., 2020b) (Lu et al., 2020b) (Sunyoto et al., 2016)

Bamboo trimmings

dOH, dCOOH, dNH2

Rice husk

–

21.48–418.33, increased with pyrolysis temperature –

Algae

–

–

2000 mg COD/L

Coconut

C]C, CdH, CdC CdO

29.2

170 mg/L

3500–4300 mg COD/ L, increased with the increase of biochar contents 370 mg/L

2.3

170 mg/L

280 mg/L

0.6 g/L

–

130

27 mmol/L

18 mmol/L

8.3 g/L

Longan shell Pine sawdust

Ref.

biochars provides habitation support for microbial attachment (Lu et al., 2020b). The pore size of biochar determines whether the organisms can be hosted on this carbon-based material and resist environmental changes by forming biofilm (Wu et al., 2020b). Smaller pore size and rough surface of biochar might supply better colonization environments for attaching organisms, especially for the microbial consortium with lower growth rates. Some researchers denoted DIET among organisms and biochars as the third reason for the manipulation of biochar on VFAs production (Fig. 3) (Lee and Lee, 2019). DIET is performed with electrons being exchanged among interspecies cells over long distance (Zhao et al., 2017). Unlike traditional IET for methane production, there is no need to form H2 as mediator when organisms adopting DIET mechanisms to convey the excess electrons produced in VFAs production to methanogens (Wu et al., 2020b). Given the avoidance of H2 formation is thermodynamically favorable (Wu et al., 2020b), the digesters containing

FIG. 3 DIET between organics-oxidizing organisms and methanogens with biochar involved.

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biochar are therefore more likely to obtain higher conversion of VFAs to CH4 than the systems without biochar addition. Hence, the VFAs accumulation would be less, if DIET is the main driver to control the degradation of carboxylates. Ameliorating the inhibition induced by the ammonia nitrogen concentrations during AD is the fourth explanation for the enhancement of VFAs by biochars. The bio-availability of nitrogen is essential for the cellular metabolism of most microbiomes. Nevertheless, excessive level of free ammonia could suppress the biologic activity of bacteria due to its bio-toxicity (Eq. 3) (Qiu et al., 2019a). Given to (1) the biofilm formed within the porosity of biochar together with (2) the reaction among ammonia and the surface carboxyl group of biochar (Bailey et al., 2011; Spokas et al., 2011), the microbial inhibition resulted from high ammonia level can then be moderated after dosing biochar, leading to higher VFAs accumulation consequently. NH+4 + OH
$ NH3 + H2 O

(3)

In sum, the VFAs output from AD can indeed be enhanced by biochar dosage, improving the resource utilization of organic waste subsequently. Biochar with rich functional groups and smaller pores can act as biofilm carrier and attribute to the VFAs accumulation. The amelioration of nitrogen inhibition s is the other explanation for higher VFAs production.

4.2 Mechanisms and enhancement of hydrogen production Although H2 is a renewable energy carrier, the generation of H2 from fossil fuels could counter to the aim of carbon emissions reduction. Therefore, the idea of producing H2 biologically via AD has gained increasing attention recently (Sharma and Melkania, 2017). Biochars, the porous eco-compatible carbon rich material, were then added to anaerobic digester to improve the H2 production (Sharma and Melkania, 2017; Watanabe et al., 2013a). For instance, Sunyoto et al. (2016) added 8.3 g/L biochar to increase H2 yield by 25.8% from food waste in a two-phase AD system (Sunyoto et al., 2016). 15.9% higher biohydrogen were obtained in the digester amended with sawdust biochar (Yang and Wang, 2019). Mechanisms of H2 enhancement resulted from biochar dosage have been studied lately as reviewed below. The pH buffering effects and the special properties of biochar were suggested as the main reasons for higher H2 production. Specifically, in addition to providing support for microbial metabolism and growth, biochar is also speculated to obtain the capacity of regulating the pH condition of cultures in anaerobic digester. The functional groups related to the pH buffering effects of biochar are controlled by pyrolysis temperature. In detail, for the biochar produced at higher pyrolysis temperatures (500–700°C), carbonates are the major alkalis to regulate the pH conditions in the digester (Wang et al., 2018b). For the case of lower pyrolysis temperatures, organic functional groups including dCOOd and dOd are the major alkalinity resources (300°C) (Wang et al., 2018b). Compared to the biochar produced under low temperature, the biochar generated under high

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temperature conditions showed greater promotion efficiency of H2 yield (Wang et al., 2018b). The adsorption of free ammonia by biochar is another reason for the changing H2 yields (Salerno et al., 2006; Sunyoto et al., 2016). As described in Section 4.1, pore size is one of the key factors controlling the adsorption ability of biochar (Salerno et al., 2006). The absorption of eco-toxic ammonia by biochar is also managed by other factors including pH, temperature and contact time. However, studies regarding the optimization of these conditions to achieve better adsorption efficiency of free ammonia are still in vogue. Notably, researchers have synthesized a biochar-supported nanoscale zero-valent iron to improve its ion exchange capability and learn its effects on H2 production from AD (Fe0 + 2H+ ! Fe2+ + H2) (Zhang et al., 2019). Excessive dosage (4.0, 5.0 mg COD/L) of this synthesized material can cause serious damage to membranes, and therefore inducing less H2 and CH4 produced by hydrogen methanogens consequently (Zhang et al., 2019). The dominant type of VFAs in digester can also regulate the biohydrogen yields to some degree (Eqs. 4–7). Theoretically, it is more efficient to produce H2 via acetate-type fermentation pathway than that of butyrate-type. However, the specific alternation of H2 production caused by the variation in VFAs composition has not been clearly revealed. Given biochar addition can also affect VFAs output biologically, more efforts are required to solve this issue. C6 H12 O6 + 2H2 O ! 2CH3 COOH + CO2 + 4H2

(4)

C6 H12 O6 ! CH3 CH2 CH2 COOH + 2CO2 + 2H2

(5)

C6 H12 O6 + 2H2 ! 2CH3 CH2 COOH + 2H2 O

(6)

2CO2 + 4H2 ! 2CH3 COOH + CO2 + 4H2

(7)

In short, biochar supplementation can improve the proportion of H2 in biogas by acting as pH buffer, mitigating ammonia inhibition, and changing the VFAs fermentation pathway. Except understanding the specific impacts of the said reasons, future studies should also centralize on investigating the eco-toxic effects of trace elements (e.g., heavy metals) contained in biochars on H2 production. Additionally, although researchers have put great efforts in investigating the promotion of H2 caused by biochar addition under anaerobic conditions, most of the studies mainly centered on the lignocellulosic wastes rather than sewage sludge (Zhao et al., 2020; Zhao et al., 2021a). Further efforts are still warranted to learn the mechanisms and enhancement of H2 via digesting anaerobic sludge in detail.

5. Sustainable application of biochar in improving aerobic sludge treatment To promote eco-friendly recycling of sewage sludge, such organic waste needs to be managed by revalorization. Anaerobic composting is a biological process in which organic matter are converted into relatively stable humus-like materials while lowering the

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245

moisture content in the sludge to achieve sludge reduction. However, composting sewage sludge directly can be hindered by the eco-toxic component (e.g., heavy metals) containing in sludge (Malinska et al., 2017). Therefore, the addition of supplementary materials into the sludge is one common strategy to solve this issue. Biochar has been considered as a potential amendment in composting various organic wastes including sewage sludge. This section will present the effects of biochar on sludge composting from the aspects of latent mechanisms, sludge reduction performance and the changes in microbial community when conducting this biological process.

5.1 Interaction mechanisms Huge quantity of sewage sludge produced from wastewater treatment has become a serious environmental issue (Cai et al., 2016). Composting is one acceptable and economically method for recycling sludge, as the organic wastes in aerobic sludge can be transformed into fairly stable humus-like product which can be adopted as soil amendment or organic fertilizer later on (Oleszczuk et al., 2014; Wu et al., 2017c). Additionally, composting is also an effective way to achieve sludge reduction, since the moisture content in sludge could be decreased dramatically after days of composting (Du et al., 2019a). However, some drawbacks associated with the composting of sludge could hinder the application of this method, including nitrogen loss (Eq. 8), unavoidable greenhouse gas (GHG) emissions, and insufficient humification of organic matter (Awasthi et al., 2016). Due to the unique properties of biochar including chemical recalcitrance, good sorption capacity resulted from its high porosity, and large surface area obtained during pyrolysis, this coal-like material has been used as an amendment to facilitate the composting of aerobic sludge. NH+4 ðcompostÞ $ NH3 ðcompostÞ + H+ $ NH3 ðgasÞ $ NH3 ðatmosphereÞ

(8)

Biochar addition is an applicable strategy for avoiding N loss especially when composting organic wastes with high N content or low C/N (Waszkielis et al., 2013). Up to 64.1% less nitrogen loss was attained when incorporated 9% of biochar with sludge (Hua et al., 2009). The high sorption capacity and large surface area of biochar ensured high NH+4 adsorption and thus reduce the N loss (Hua et al., 2009). The nitrifying bacteria immobilizing on the char is another reason for lower N loss in the biochar-amended system, as which could transfer NH+4 into NO
3 biologically (Zhang and Sun, 2014). GHG driven from compost are generally CO2, CH4, and nitrous oxide (N2O). Compared to the other two GHG, N2O is the most unwanted gas as which has the highest global warming potential (265-fold greater than CO2) (Wu et al., 2020a). Researchers have found the 5%–10% biochar dosage can efficiently decrease N2O emissions by up to 30%–40% (Z˙ukowska et al., 2019). Moreover, since N2O emission is partially responsible for the nitrogen loss, mitigating the N2O generation can in turn prevent the ammonia volatilization to some degree. Considering N2O can also be generated concurrently by denitrifying organisms with CH4 under anaerobic conditions, one additional benefits of biochar dosage is the boost of CH4 yield while repressing N2O production in the absence of O2 (Fig. 4).

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FIG. 4 Conceptual image of the biochar addition to sludge treatment and associated effects.

Given the amount of humic substances is important to soil ecology which can impact plant growth thereafter, it is essential to enhance the humification of sludge and convert organic matter into more stable humic substances. Biochar has been adopted to attain that objective, as the humification of sludge is controlled by the type and quality of bulking agent (Goyal et al., 2005). For instance, up to 30% of water soluble C was converted into humic substances after biochar addition ( Jindo et al., 2012; Song and Guo, 2012). The reactions among the surface oxygen functional groups on biochar and the soluble organic C in sludge may explain that high conversion efficiency ( Jindo et al., 2012, Song and Guo, 2012). Alternations in the surface properties of biochar such as the increase in carboxylic group were attained when combining biochar with composting together (Wu et al., 2017b). The oxidation of the biochar surfaces itself or the sorption of humic acid is the possible reason for such alternation. No studies have compared the effect of original biochar and the biochar with the said alternation on sludge humification. Further studies are warranted to understand this issue. Collectively, biochar addition is a relatively simple and feasible strategy to assist the composting of sewage sludge by easing associated greenhouse effect, reducing nitrogen loss and improving humification. Despite the positive effects of biochar on sludge composting has been clearly documented, our current understanding on the specific mechanisms of the interactions among biochar and sludge are still not enough. Studies regarding the effects of humic acid adsorption on biochar structure and following sludge composting are still warranted.

5.2 Sludge reduction performance Next to temperature, moisture is another important index in anaerobic composting. This is because such parameter indicates the sludge volume and reflects the effectiveness of the process of composting. During the process of composting, the alternation in moisture content is primarily induced by the degradation of organic waste and the evaporation caused by high temperature or ventilation (Liu et al., 2021). Seeking suitable moisture level is important, as too high or too low moisture of composted raw material can suppress oxygen flow within the composted piles and reduce microbial activity, respectively (Godlewska et al., 2017). Compared to the amount of water generated from organic matter

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degradation, the water removed by evaporation and ventilation via heat is greater, decreasing the moisture content of composting piles thereafter. Moisture content of composting process can be optimized thanks to the addition of biochar. The moisture content could be decreased from initial 70%–80% to 50%–60% (Liu et al., 2021). For instance, a decrement in moisture from 74.43% to 68.67% were also been found when studying the effects of biochar amendment on vermicomposts (Malinska et al., 2017). Biochar particle size was reported to influence sludge reduction via composting (Huet et al., 2012). Bigger particle size of bulking agent could augment the air permeability, inducing higher heat transfer efficiency in composting (Du et al., 2019a). However, big biochar cannot kill pathogens effectively during the composting of sewage sludge, since which is not good at insulating and heating of compost materials (Du et al., 2019a). Therefore, applying big biochar to composting are not likely to induce lower moisture content compared to small one.

5.3 Microbial community Composting is a biological process controlled by microbes and associated enzymatic activities. Environmental conditions including pH, temperature, moisture content and nutrient level can influence the microbial community structure. In an attempt to regulate the composting process with biochar addition effectively, understanding the composition of reactor microbes and their change is necessary. Despite the microbial community in composting has been studied for decades, the traditional molecular tools cannot provide a comprehensive evaluation of the microbial diversity and structure (Ogino et al., 2001). Advanced molecular technologies grounded on 16S rDNA sequence technologies have been proven as an effective tool to reveal the microbial dynamics during sludge composting (Awasthi et al., 2017b). Awasthi et al. (2016) are the researchers using aforementioned advanced molecular tools and found that the bacterial diversity was negatively related with biochar dosage (Awasthi et al., 2017b). Proteobacteria, Firmicutes, and Chloroflexi were the dominant phyla in the sewage sludge composting amended with different biochar (Awasthi et al., 2017b). Clostridium and Lactobacillus, the main genera of Firmicutes, were also found during sludge composting. These two kinds of bacteria are usually observed under high levels of organic acids and ethanol (Gibello et al., 2011; Wakase et al., 2008). Genera Paracoccus, Planomicrobium, Devosia, and Agrobacterium were deemed to be closely linked with main composting factors (Awasthi et al., 2017b). Therefore, the change of these genera can act as the indicator to regulate the composting conditions. Except bacteria, fungi were also detected during the cocomposting of high ammonia-contained sludge mixed with organic fraction of municipal solid waste under thermophilic conditions (Awasthi et al., 2017a). Ascomycota and Basidiomycota were reported as the dominant fungal phylum in this composting system (Awasthi et al., 2017a). Most of the studies related to the microbial dynamics when conducting composting process were mainly focused on heterotrophic microbial consortium and their enzymatic activities. However, the

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understanding of the changes in fungal diversity is not as sufficient as their counterparts, especially when composting sludge with biochar supplied. Considering fungi can not only degrade complex organic matter into easily degradable substances but also obtain the capability to secrete various extracellular enzymes to facilitate composting process, ongoing investigation regarding the fungal dynamics during composting is crucial (Wu et al., 2017a). Similar to Awasthi’s study, Du et al. (2019a, b) also found that biochar addition can change the functional organism community and induced following impacts on composting results (Du et al., 2019a, b). The activities of most enzymes and functional microbial communities can be strengthened by biochar addition. Specifically, although the abundance of some composting organisms such as Trichoccus and Saccharibacteria were negatively correlated with biochar addition, a positive relationship among biochar addition and other functional bacteria including Bacillus and Saprospiraceae were still detected in Du’s studies. Temperature is a principle factor affecting the microbial community during sewage sludge composting (Koyama et al., 2018). It not merely influences the rate of composting but also eliminates the threat posed by pathogens present in organic wastes originating from living microorganisms. Shifts in microbial community were found when increasing temperature from 50°C to 70°C (Koyama et al., 2018). Biochar dosage could enhance aeration rate and therefore improve the proliferation of organisms, accelerating the degradation of organic wastes consequently (Zhang et al., 2016). Moreover, biochar, especially at higher dosage (10%–20%), could weaken the relationship between temperature and functional bacterial community, but strengthen the link between enzymatic activities and functional bacterial communities in Du’s study (Du et al., 2019a, b). In detail, the activity of dehydrogenase is closely linked with Saprospiraceae abundance. Researchers attributed the higher cellulose activity detected in composting piles with biochar addition to the important roles of Bacillus in cellulose secretion (Du et al., 2019b). Other enzymes responsible for organic matter degradation or hydrolysis (i.e., protease and urease) were all reported to show different patterns of changes in activities (Du et al., 2019a). However, these two enzymes did not exhibit any correlation with the abundance of organisms in the composting piles (Du et al., 2019a). Overall, biochar is a potential amendment in improving final compost quality by alternating microbial community and changing associated enzymatic activities. Correlations among some functional organisms, the associated enzymatic activities, and biochar addition have been reported lately. However, the relationship between the physicochemical properties of biochar and the shift in microbial communities has not been clearly revealed and require more experimental investigation.

6. Conclusions and perspectives Biochar could significantly encourage sludge dewaterability. This improvement is mainly attributed to the generation of free radicals from the decomposition of oxidants by

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biochar. Specifically, as such type of water combined with EPS cannot be removed by the conventional dewatering process, the biochar-derived free radicals would than destroy the sludge floc and EPS to further release the interstitial water existing in the sludge flocs, achieving better dewatering thereafter. In addition, the permeable and rigid lattice structure of biochar allows such carbon-rich material to ensure an unblocked water filtration channel and to further promote dewaterability. Although biochar/peroxide exhibits attractive effects on sludge dewatering, there are still some deficiencies regarding this impact that should be addressed in future research. Adjusting pH level to the optimum range is necessary for the best performance of biochar/H2O2 system on dewatering. Despite replacing H2O2 with persulfate could expand this pH range, the transformation of sulfate ion has a negative effect on follow-up sludge treatment. In addition, some metal ions existing in the sludge may react with free radicals that should be used to oxidize the EPS, which drops the sludge dewaterability. Thus, it is necessary for further improving the performance to address these issues. In AD, biochar also has positive effects on methane production. The functions of biochar in facilitating AD are multiple, including acting as adsorption of inhibitors, pH buffer, promoter of electron transfer efficiency and the microbial carrier for functional organism enrichment. DIET is the potential explanation for encouraging methane production. Current evidence of DIET in biochar-based AD is provided indirectly by searching known bacterial and archaeal species with DIET function. Some persistent barriers should be overcome to expand the application of biochar amended AD processes. The inhibitory effect on AD process observer by some researchers should be totally releveled in the future. Seeking direct evidence is required to confirm enhanced IET by biochar from the view of microbial metabolic mechanisms. A technical and economic evaluation regarding effects of biochar addition on AD is also required to scale-up the application of biochar in this biological process. Biochar has been used as additive/support media to influence VFA and H2 production through fermenting sludge. Alternations in VFA production resulted from adding this carbon-rich material are ascribed to the shift microbial community structure, improve the immobilization of organisms with low growth rate, augment electron transfer, and mitigating the inhibition caused by ammonia nitrogen. The presence of biochar can also improve the yield of H2 from fermentation due to its ability to buffer pH conditions and ameliorating the eco-toxic of ammonia. Ongoing investigation should be more centered on learning the effectiveness of adopting biochar as amendment to facilitate sewage sludge instead of lignocellulosic wastes. Studies concerning the impact of trace elements containing in the biochar on H2 yield during sludge digestion has not been thoroughly examined. Intensive efforts regarding this issue are therefore required. Composting is a self-heating biological process in which organic wastes convert into humus-like substances by indigenous organisms including bacteria and fungi. Biochar is a potential amendment to accelerate composting process and improve associated final compost quality by stimulating microbial and enzymatic activity. Aerobic sludge can be dramatically reduced via composting after biochar addition. Small biochar particle type

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favors sludge reduction over bigger one as which is better at insulating and heating compost matters. More efforts are required to study the interactions among biochar dosage and the organisms involved in microbial humification, an important process to transform easily biodegradable organic waste into humic substances. Further evaluation regarding the correlation among the changes in physicochemical characteristics of biochar and associated microbial alternation is deemed necessary.

Acknowledgments This work is supported by an Australian Research Council (ARC) Future Fellowship (FT160100195). Ms. Lan Wu is supported by China Scholarship Council (File No. 201806330112) for a PhD scholarship at the University of Technology Sydney.

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9 Sustainable application of biochar for storm water reuse Van Son Trana, Huu Hao Ngob, Wenshan Guob, and Manh Khai Nguyena FACULTY O F E NVIR ONMENTAL SCI ENCES, U N IV ER SIT Y O F SC IENCE, VIETNAM NATIONAL UNIVERSITY, HANOI, VIETNAM b CENT RE F OR TEC HNO LO GY IN W ATER AND W AS TEW ATER , S C H O OL OF C I V I L A N D E NV I R ON M E N T AL EN G INEERING, UNIVERSITY OF TECHNO LOGY SYDNEY, SYDNEY, NS W, AUSTRALIA a

1. Introduction Currently, more than 53% (3.8 billion) of the world’s estimated 7.2 billion people live in cities, and this total is expected to rise to 66% by 2050, given the current unchecked rate of population growth which puts pressure on the planet’s finite resources. The growing lack of clean and sustainable water supplies has serious implications for the growing urban and rural populations (Goonetilleke et al., 2017). Seasonal and sometimes perennial water shortages are becoming a common feature in many parts of the world. This has resulted in unsustainable extraction of surface and groundwater resources in metropolitan areas, as well as deteriorating water quality. Changes in rainfall patterns and growing unpredictability of rainy seasons have accelerated the rate of depletion (Strauch et al., 2015; Yaduvanshi and Ranade, 2015). The increase in human demand and major weather changes due to climate change are exacerbating the scarcity of clean water. Although there are many options for desalination filtration or water transport between basins that seem interesting, this option is expensive and not suitable, currently, for less developed countries. In this situation, storm water is considered to be a widely untapped resource. Storm water is gaining popularity as a reusable alternative water supply, particularly in water-scarce areas. In this context, urban road storm water has great potential to be reused to mitigate water scarcity issues (Hong et al., 2018; Mohanty et al., 2018). Nevertheless, surface runoff in cemented urban areas often contains various pollutants such as chromium (Cr), copper (Cu), cadmium (Cd), lead (Pb), nickel (Ni), and zinc (Zn) (Sun et al., 2020). Storm water pollution has received a lot of attention in recent times since it can affect the viability and quality of waters in receiving water bodies. There are serious negative consequences for the environment to consider (Liu et al., 2018). With increasing urbanization, urban storm water runoff has garnered much attention due to its impact on water quality. Urban runoff contributes total suspended solids (7.8–5700 mg/L),

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nitrogen (0.4–20 mg/L), metals (160–914 μg/L), polycyclic aromatic hydrocarbons (677–6477 ng/L), phenols (400–9690 ng/L), pesticides (0.26–247 μg/L), and other contaminants to natural water bodies. For this reason, it will probably require treatment before discharge (Tian et al., 2016). Previous storm water studies have mostly concentrated on the relationship between pollution deposition on ground surfaces like roads and entrainment in storm water runoff (Wijesiri et al., 2018). Hazardous heavy metals can be carried away after a rainstorm and end up in storm water runoff. This might jeopardize the safety of storm water reuse, which is typically associated with severe environmental hazards (Hong et al., 2018). Concentrations of heavy metals are larger in industrial areas than in residential ones. Concentrations of heavy metals in storm water runoff are highly variable depending on the sampling location. The ecological risks posed in storm water runoff vary spatially, and to ensure the safety of rainwater reuse, it is important to identify areas that can be prioritized so that rainwater collection can be targeted (Liu et al., 2016). Researchers have noted that even volatile pollutants such as benzene series pollutants (BTEX) which include benzene, ethylbenzene, toluene and xylene can be washed-off by storm water runoff, resulting in storm water pollution (Mahbub et al., 2011a). In earlier studies, storm water runoff samples are often collected from catchment surfaces, in an effort to correlate their concentrations/loads with the surrounding environment (see Liu et al., 2018). Storm water quality depends not only on pollutants contributed by the ground phase, but also pollutants contributed by the atmospheric phase (air) (Weerasundara et al., 2018). Pollutants generated in the urban environment can firstly accumulate in the air and are then transported to ground surfaces through atmospheric deposition. Atmospheric deposition can occur as dry and wet deposition. In this context, atmospheric pollution can emerge as a significant contributor to storm water quality degradation. For this reason, it is crucial to find an environmentally friendly technology that is more practical and effective for application to storm water reuse. Existing physical and biological technologies for storm water reuse such as filtration, settling within storm water ponds, wetlands, biofiltration, swales, porous pavements, and green rooves could effectively remove sediments and associated pollutants, for instance, heavy metals (Payne et al., 2019). However, a large area is usually required for application of the methods. Furthermore, it is also critical to remove nontraditional pollutants in storm water for examples, herbicides, pesticides, polycyclic aromatic hydrocarbons (PAHs) and pathogens. Biochar, a posteritied transformation of lignocellulosic biomass, is often formed as a byproduct from waste biomass such as wood chips, agricultural wastes by high temperature pyrolysis under oxygen limited conditions. Biochar with 100 uses is considered as black gold (BG) of ecosystem. It is projected to survive for decades in the environment due to the refractory carbon in biochar, which has a half-life of over 100 years (Mohanty et al., 2018). The biochar properties (particle size distribution, density, CEC,

Chapter 9 • Sustainable application of biochar for storm water reuse 261

and surface area) are affected by feedstock properties and pyrolysis temperature (Rahman et al., 2020; Tran et al., 2020). Due to its low cost and high adsorption capacity, biochar added to sand biofilters has been studied for its ability to remove pollutants such as nitrogen, E. coli, heavy metals, and organic pollutants and soil amendments (Tran et al., 2020).

2. Sources, quantity and quality of storm water, reuse requirement and biomass 2.1 Sources, quantity and quality of storm water According to prior studies, storm water contains high concentrations of heavy metals (Sun et al., 2020). High quantities of zinc in urban storm water runoff degrade the quality of recipient water bodies. The researchers discovered a link between zinc concentrations in urban runoff in sediments from different locations throughout the world and the amount of fuel sold in those cities, which may be used as a proxy for traffic volume. Zinc levels in urban rainfall runoff are high, which greatly endangers the required quality of water courses. In many urban areas with much storm water runoff, zinc levels in surface waters surpass the water quality criterion (120 g/L). Significant quantities of nickel, vanadium, cadmium, and zinc enter the natural environment along with combustion of liquid fuel, diesel, and condensed hydrocarbonates. Cadmium originates from various products of type and road surfaces and it is used to coat steel objects. The most frequent sources of deposited cadmium are: metallurgical industry, chemical industry, battery manufacturers, and municipal wastewater. Nickel occurs in vehicle fuel and its presence in storm water runoff can result in ground erosion. In the last few years, emissions of arsenic into the atmosphere have risen, mainly from the combustion of crude oil and black coal (Milik and Pasela, 2018). Sediments from storm water runoff treatment devices, particularly those located in city centers, industrial districts, and near roads, contain a high level of heavy metals and been shown to be the most polluted areas. Storm water runoff, which is the major source of water quality problems in urban areas, deposits a large portion of these hazardous chemicals on the surface of highways and urban roadways, eventually depositing them into local water bodies (Milik and Pasela, 2018). Cr 128 g, Co 12.35 g, Ni 98.5 g, Cu 607.5 g, Zn 8429.5 g, As 6.95 g, Cd 3.7 g, and Pb 251.75 g were the average released quantities of heavy metals during one rainy event, demonstrating that metal runoff loads in storm water runoffs are strongly connected to nearby industries (Lee et al., 2020a). The pH of storm water is also a problem regarding the reuse of this source of water. It is reported that average pH values of storm water in Vietnam were 5.83
0.62. In some case, the pH of rainwater was less than 5.6 even those being neutralized the acid density by containing some base ions (Han et al., 2021). The low pH value prevents rainwater from being reused for purposes such as drinking and irrigation and leads to soil acidity

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Current Developments in Biotechnology and Bioengineering

Table 1 runoff.

The annual concentration of heavy metals in sediment from storm water Heavy metal concentration (mg/kg)

Country

Samples

As

Cr

Zn

Pb

Cd

Ni

Cu

Poland, Białystoka

Street inlets

3.20– 20.20

28.10– 429.00

Poland, Częstochowaa Canada, Ontarioa

Urban rainwater runoff Settling tanks and separators Industrial areas

2.50– 3.00 12.10– 28.24 4.00– 9.00

290.00– 360.00 129.00– 640.00 0.284– 6.200 640.00– 809.00 11.704– 14.029 3.290– 3.605

0.06– 39.90 0.47– 0.77 1.40– 1.90

3.10– 30.60 16.05– 32.12 21.00– 26.00

2.50– 44.60 3.37–7.99

41.00– 74.00

28.10– 429.00 21.00– 63.00 67.00– 73.00 70.00– 490.00 0.003– 0.006– 85.00– 323.00 1.003– 840.00 1.438– 706.00

0.40– 8.30 13.00– 18.00 1.90– 3.50

0.003– 0.010 32.00– 54.00 125.00– 195.00 231.00– 312.00

China, Guangzhoua Serbia, Belgradea France, Lyona Korea, Jeongwangcheonb Korea, Ansanc a b c

Urban rainwater runoff Retention reservoirs 6.60– 13.00 Industrial complex 20.00– 28.00 Industrial complex 19.00– 24.00

0.074– 1.350 40.00– 77.00 328.00– 370.00 579.00– 769.00

71.00– 98.00 24.00– 201.00 0.067– 1.820 130.00– 349.00 5.506– 6.114 1.810– 1.997

Milik and Pasela (2018). Lee et al. (2020a). Lee et al. (2020b).

if rainwater enters the soil. In their study, Chintala et al. (2014) confirmed that application of biochar could significantly increase the soil pH and decreased the exchangeable acidity (Table 1). Furthermore, past researchers did note that even volatile pollutants such as benzene series pollutants or volatile organic compounds (VOCs) which include benzene, ethylbenzene, toluene and xylene, could accumulate on road surfaces (Mahbub et al., 2011b). Storm water runoff can wash VOCs contaminants away, and this results in storm water contamination (Mahbub et al., 2011b) (Table 2).

2.2 Requirements for storm water reuse The notion of “water suited for purpose” should guide storm water reuse. This requires using water of varying quality depending on the intended application. This enables water of varied quality to also help create less expensive water treatment, with only the best grade water being utilized for direct consumption. This subsequently necessitates a variety of already-existing suitable technologies, yet it does not require the creation of new technologies. Instead what is needed is to refine current technologies and application techniques to resolve certain scenarios (Goonetilleke et al., 2017). Table 3 describes the treatment purposes and required water quality for storm water reuse.

Chapter 9 • Sustainable application of biochar for storm water reuse 263

Table 2

The annual concentration of VOCs from storm water runoff.

Country

Pollutant

Unit

Concentration

References

Shenzhen China Urban road

Toluene Ethylbenzene m- and p-xylene o-xylene Methyl-tert-butyl ether Benzene Toluene Ethylbenzene m-xylene o-xylene p-xylene 1,2,4-Trimethylbenzene Styrene Naphthalene Chloroform Bromodichloromethane Vinyl chloride Methyl tert-butyl ether Benzene Toluene Ethylbenzene m-xylene o-xylene p-xylene 1,2,4-Trimethylbenzene Styrene Naphthalene Chloroform Vinyl chloride Methyl tert-butyl ether Benzene Toluene Ethylbenzene m-xylene o-xylene p-xylene 1,2,4-Trimethylbenzene Styrene Naphthalene Chloroform Vinyl chloride

μg/m2

0.189–0.757 0.089–0.413 0.167–0.691 0.064–0.259 0.024
0.008 0.040
0.022 0.042
0.011 0.590
0.004 0.622
0.006 0.018
0.006 0.561
0.005 0.012
0.005 0.035
0.016 0.062
0.020, 0.018
0.006 0.036
0.023 0.794
0.032 0.183
0.105 0.032
0.007 0.035
0.006 0.586
0.001 0.617
0.002 0.013
0.002 0.557
0.001 0.008
0.002 0.016
0.003 0.038
0.010 0.024
0.004 0.787
0.061 0.032
0.025 0.054
0.011 0.033
0.004 0.587
0.002 0.621
0.003 0.015
0.001 0.560
0.007 0.009
0.002 0.048
0.015 0.054
0.013 0.024
0.003 0.839
0.041

Hong et al. (2018)

Beijing, China

Highway junction

Gas station

City road

μg/L (average
SD)

μg/L (average
SD)

μg/L (average
SD)

Li et al. (2018)

Li et al. (2018)

Li et al. (2018)

Continued

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Current Developments in Biotechnology and Bioengineering

Table 2

The annual concentration of VOCs from storm water runoff—cont’d

Country University campus

Park

Residential

Seoul, Korea

Urban drain

North Carolina, Campus USA

Pollutant

Unit

Concentration

References

Methyl tert-butyl ether Benzene Toluene Ethylbenzene m-xylene o-xylene p-xylene 1,2,4-Trimethylbenzene Styrene Naphthalene Chloroform Vinyl chloride Methyl tert-butyl ether Benzene Toluene Ethylbenzene m-xylene o-xylene p-xylene 1,2,4-Trimethylbenzene Styrene Naphthalene Chloroform Vinyl chloride Methyl tert-butyl ether Benzene Toluene Ethylbenzene m-xylene o-xylene p-xylene 1,2,4-Trimethylbenzene Styrene Naphthalene Chloroform Vinyl chloride Methyl tert-butyl ether Benzene Toluene m- and p-xylene Benzene Toluene m- and p-xylene

μg/L (average
SD)

0.038
0.013 0.033
0.009 0.026
0.005 0.584
0.001 0.613
0.002 0.011
0.001 0.557
0.003 0.006
0.002 0.033
0.032 0.044
0.006 0.025
0.003 0.903
0.090 0.025
0.010 0.030
0.008 0.021
0.004 0.584
0.001 0.613
0.004 0.010
0.002 0.559
0.003 0.004
0.001 0.015
0.009 0.034
0.005 0.026
0.015 0.827
0.043 0.018
0.010 0.026
0.013 0.020
0.002 0.588
0.006 0.617
0.006 0.010
0.002 0.561
0.006 0.007
0.002 0.022
0.009 0.042
0.008 0.024
0.009 0.833
0.020 1.58 1.71 822.53 334.00 0.04
0.05 0.06
0.10 0.02
0.03

Li et al. (2018)

μg/L (average
SD)

μg/L (average
SD)

μg/L

nM

Li et al. (2018)

Li et al. (2018)

Yu et al. (2017a)

Mullaugh et al. (2015)

Chapter 9 • Sustainable application of biochar for storm water reuse 265

Table 3 Summary of treatment purposes and required water quality for storm water reuse in the United States (Luthy et al., 2019). Area

Purpose

Water quality requirements

State of Minnesota, USA

Unrestricted irrigation

District of Columbia

Unrestricted irrigation

Turbidity 3 NTU TSS 5 mg/L pH 6–9 Chloride 500 mg Zinc 10 mg/L Copper 5 mg/L Pathogens/indicators: E. coli: 126 CFU/100 mL Zinc 15 mg/L Pathogens/indicators: E. coli: 4615 CFU/100 mL, Crypto.: 0.033 oocysts/L Zinc 160 mg/L Pathogens/indicators: E. coli: 50,000 CFU/100 mL, Crypto.: 0.320 oocysts/L BOD5 10 mg/L Turbidity 2 NTU TSS 10 mg/L pH 6–9 Pathogens/indicators: E. coli: 2.2 CFU/100 mL Turbidity 2 NTU Pathogens/indicators: Virus: 3.0-log reduction, Bacteria: 2.0-log reduction

Indoor use

Los Angeles, CA

Unrestricted irrigation, indoor use

San Francisco, CA

Unrestricted irrigation, indoor use

3. Biomass and modifying technique for biochar production 3.1 Biomass for biochar preparation Huge amounts of cheap biomass belonging to some principal categories such as lignocellulose, algae, chitin/chitosan, activated sludge, bacteria biomass, fungal biomass, etc., can serve as adsorbents for heavy metals and removal of organic pollutants (Tran et al., 2015). Among these categories, lignocellulose could be employed for biochar preparation. Chemical composition, functional groups, surface area, porosity, and surface morphology are important factor to determine pollutants abatement on lignocellulosic materials. Fig. 1 showed chemical compositions of some typical low-cost lignocellulosic adsorbents. Plant residues such as wooden materials, coconut shell, pineapple leaves, banana stem, sugar cane bagasse, and coffee waste have the largest amounts of cellulose (>40%). The highest lignin content (>30%) appears in biomass sources like neptune grass, soft wood, and coconut shell and bark (Fig. 1). It has been reported that fruits and vegetables accounted

266

Current Developments in Biotechnology and Bioengineering

100 90

others ash

80

C h e m i c a l c o m p o si t i o n ( % )

Lignin 70 60 50

Hemicellulose Cellulose

40 30 20 10 0

FIG. 1 Average chemical composition of some typical lignocellulose (% dry weight) (Tran et al., 2015).

for almost 50% of food waste generated by European households (De Laurentiis et al., 2018). Removal capacities of biomass materials could be affected by some factors such as polarity and aromaticity (Xi and Chen, 2014) and lignin is mainly responsible for abatement of organic pollutants as well as various pretreatment and/or fabrication methods (e.g., Nguyen et al., 2019; Hoang et al., 2020; Tran et al., 2020). In addition, ash content of lignocellulose also played an important role in the adsorption process of organic pollutants on these materials. The existence of functional groups such as hydroxyl (–OH), carboxyl (–COOH) and silanol (Si-OH) is also an important factor for sorption of organic pollutants, for instance, pesticides, PAHs, from water (Abdolali et al., 2014; Tran et al., 2015). The surface area and porosity of lignocellulose also greatly affect their pollutants adsorption capacities. The adsorbents have higher surface area and then greater adsorption capacities for contaminants (Tran et al., 2015).

3.2 Modifying technique for biochar production Many studies utilized the lignocellulosic waste and by-products to remove metal ions and the Langmuir maximum adsorption capacities of heavy metals onto the adsorbents are indicated in Table 4. Biomass modification for removing antibiotics from water has been researched by several investigators. Yu et al. (2017b) evaluated the viability of levofloxacin removal from

Chapter 9 • Sustainable application of biochar for storm water reuse 267

Table 4 Modifying techniques for biochar production and other adsorbents from biomass to remove pollutants. Pollutants

Sorbents (modifying agent)

Qmax (mg/g)

References

Pb

Rice husk (alkali and heat) Rice husk (tartaric acid) Sawdust (formaldehyde) Natural seaweed

58 108 37.0 369.6

Zn

Rice husk (alkali and heat) Carrot residues (HCl) Natural seaweed

8.1 29.6 46.3

Cd

Rice husk (alkali and heat) Green coconut shell powder Natural seaweed

16.7 285.7 95.8

Cu

Rice husk (alkali and heat) Rice husk (tartaric acid) Sawdust (formaldehyde) Carrot residues (HCl) Seaweed (2-propanolits) Natural seaweed

10.9 29 37.2 32.7 143.3 109.9

Ni

Rice husk (alkali and heat) Seaweed (2-propanolits) Natural seaweed

5.5 131.6 55.5

As

Rice polish

As3+138.88; As5+ 147.05 As3+ 128.10 As3+ 704.11 36.1 35.71 35.71 45.1

Krishnani et al. (2008) Wong et al. (2003) Ahmad et al. (2009) Ahmady-Asbchin et al. (2009) Krishnani et al. (2008) Nasernejad et al. (2005) Ahmady-Asbchin et al. (2009) Krishnani et al. (2008) Pino et al. (2006) Ahmady-Asbchin et al. (2009) Krishnani et al. (2008) Wong et al. (2003) Ahmad et al. (2009) Nasernejad et al. (2005) Basha et al. (2009) Ahmady-Asbchin et al. (2009) Krishnani et al. (2008) Basha et al. (2009) Ahmady-Asbchin et al. (2009) Ranjan et al. (2009)

Heavy metals

Hg

Fruit rinds of G. cambogia (fresh) Fruit rinds of G. cambogia (immobilized) Rice husk (alkali and heat) Sugarcane bagasse (raw) Sugarcane bagasse Carrot residues (HCl)

Kamala et al. (2005) Kamala et al. (2005) Krishnani et al. (2008) Abdolali et al. (2014) Khoramzadeh et al. (2013) Nasernejad et al. (2005)

Antibiotics Tetracycline Tetracycline Tetracycline Levofloxacin Ciprofloxacin Norfloxacin Metronidazole Phenol

Biochar derived from fast pyrolysis of biomass Activated carbon from beet pulp Activated carbon from peanut hulls Zirconium (IV)-loaded corn bracts Microporous activated carbon from lignocellulosic biomass by microwave pyrolysis Microporous activated carbon from lignocellulosic biomass by microwave pyrolysis Acid and thermal modified canola residues

58.8 288 28 73 131.14

Rice-straw-based carbon Rice husk-activated carbon

14.2 27.58

166.99 21.42

Liu et al. (2012) Torres-Pe
rez et al. (2012) Torres-P
erez et al. (2012) Yu et al. (2017b) Ahmed and Theydan (2014) Ahmed and Theydan (2014) Balarak and Mostafapour (2016) Wang et al. (2007) Kalderis et al. (2008) Continued

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Current Developments in Biotechnology and Bioengineering

Table 4 Modifying techniques for biochar production and other adsorbents from biomass to remove pollutants—cont’d Pollutants

Sorbents (modifying agent)

Qmax (mg/g)

References

Banana peel (raw) Corn grain-based activated carbon

688.9 256

Achak et al. (2009) Park et al. (2010)

Pine wood (raw) Pine wood (desugared) Pine wood (raw) Pine wood (desugared) Pine needles (raw) Pine needles (desugared) Rice husk (H3PO4, heat)

1497.9  103 7533.47  103 417.7  103 3084.4  103 447.56  103 3103.27  103 104.5 m

Benzo(a)pyrene

Cork waste (raw)

Benz(a)anthracene

Cork waste (raw)

Chrysene

Cork waste (raw)

Benzo(ghi)perylen

Cork waste (raw)

(21
5)  10– 3 m (26
8)  10– 3 m (23
5)  10– 3 m (23
2)  10– 3 m

Xi and Chen (2014) Xi and Chen (2014) Xi and Chen (2014) Xi and Chen (2014) Xi and Chen (2014) Xi and Chen (2014) Yakout and Daifullah (2013) Olivella et al. (2011)

PAHs Acenaphthene Pyrene

Olivella et al. (2011) Olivella et al. (2011) Olivella et al. (2011)

Organic pesticides DDD Dieldrin Endrin 4-Chloro-2-methyl phenoxy acetic acid Bromopropylate

4,4-DDT 4,4-DDT

Bagasse fly ash Olive stone (acid-treated) Olive stone (acid-treated) Coffee wastes

7.69  103 23.74 43.71 340

Activated carbon from corn cob Activated carbon from olive kernel Activated carbon from soya stalks Activated carbon from rapeseed stalks Commercial activated carbon (Filtrasorb400, Calgon company) Commercial activated carbon (NORIT® GL50) Wood sawdust Commercial activated carbon

18.9  102 12.3  102 11.6  102 7.9  102 15.6  102

Zolgharnein et al. (2011) Zolgharnein et al. (2011) Zolgharnein et al. (2011) Al-Zaben and Mekhamer (2017) Ioannidou et al. (2010) Ioannidou et al. (2010) Ioannidou et al. (2010) Ioannidou et al. (2010) Ioannidou et al. (2010)

21.17  102 4.25–69.44 163.9

Ioannidou et al. (2010) Boussahel et al. (2009) Boussahel et al. (2009)

Corn cob Lignocellulosic substrate Lignocellulosic substrate Vegetable activated carbon Coconut activated carbon Vegetable activated carbon Coconut activated carbon Vegetable activated carbon Coconut activated carbon

4.07 61.8 11.2 23.51 26.02 25.62 27.15 23.92 26.32

Ara et al. (2013) Boudesocque et al. (2008) Boudesocque et al. (2008) Moreno et al. (2010) Moreno et al. (2010) Moreno et al. (2010) Moreno et al. (2010) Moreno et al. (2010) Moreno et al. (2010)

Organic herbicides Metribuzin Isoproturon Terbumeton Prometon Propazine Prometryn

Chapter 9 • Sustainable application of biochar for storm water reuse 269

water using zirconium (IV) loaded corn bracts. The adsorbent exhibited a strong adsorption capacity (Qmax ¼ 73 mg/g) and it could also regenerate with a desorption rate as high as 89% with pH 11. In another study, removal efficiencies of fluoroquinolones antibiotics by microporous activated carbon from lignocellulosic biomass by microwave pyrolysis were 96.12% of ciprofloxacin and 98.13% of norfloxacin, respectively (Ahmed and Theydan, 2014). Maximum Langmuir adsorption capacities of 131.14 and 166.99 mg/g for ciprofloxacin and norfloxacin on the adsorbent were achieved, respectively. Balarak and Mostafapour (2016) treated canola residues at a temperature of 105°C and 0.5 M H2SO4 as an adsorbent to remove metronidazole antibiotic. The theoretical maximum of adsorption capacity following the Langmuir model was 21.42 mg/g, and the equilibrium status was achieved within 90 min of contact time. Some authors have utilized activated carbon obtained from lignocellulose to take up antibiotics from aqueous solution. For example, Ilhan et al. (2012) used woodchip for removing antibiotics from water and concluded that 65% of the sorbed atrazine, 70% of sulfamethazine, 90% of enrofloxacin, and 80% of monensin A were reserved in wood chips. In another study, removal of antibiotics by coagulation and granular activated carbon filtration was done (Choi et al., 2008). While removal efficiencies of oxytetracycline-HCl, democlocycline-HCl, and tetracycline were more than 90%, less than 70% of
rez et al. minocycline-HCl and meclocycline-sulfosalicylate were removed. Torres-Pe (2012) used activated carbons from agricultural residues to abate antibiotics from water and removal efficiencies reduced according the following order: commercial granular activated carbon GAC2 (817 mg/g) > beet pulp BP-H2O (288 mg/g) > commercial granular activated carbon GAC1 (133 mg/g) > peanut hulls PH-H2O (28 mg/g). In another study, penicillin G uptake capacity was determined to be 330.0 mg/g for activated sludge, 459.0 mg/g for R. arrhizus, and 375.0 mg/g for activated carbon at the experimental conditions (Aksu and Tunc¸, 2005). Several typical food processing waste biosorbents such as rice husk, Azolla filiculoides, apricot stone shells, tendu leaf, chitin, and chitosan were investigated for phenolic compounds removal from water and wastewater. Of the different types of biosorbents, modified rice husk ash, banana peel, apricot stone activated carbon and functional chitosan had the highest adsorption capacities for phenolic compounds binding. Removing certain types of pesticides such as ethoprophos, oxamyl, DDT and its derivatives, i.e., lindane, malathion, aldrin, dieldrin, endrin, 4-chloro-2-methyl phenoxy acetic acid, quinalphos, α-cypermethrin, methyl parathion, quinalphos, and bromopropylate by biosorbents, has also been researched. As can be seen in Table 4, adsorption capacities of pesticides on biosorbents were often lower than commercial activated carbon. However, biosorbents can be applied to remove many different kinds of pesticides and several biosorbents have very good binding capacity such as chitosan, activated carbon from apricot stone, acid modified olive stone, coffee wastes and activated carbon from corn cob.

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Current Developments in Biotechnology and Bioengineering

4. Application of biochar for storm water reuse Treated storm water has been used for drinking and nondrinking purposes such as gardening, toilet flushing, fire protection and flood control. It is reported that rainwater has rarely been utilized for drinking purposes, especially in urban areas (only 2% in Australia) (Bui et al., 2021). Before utilization for different purposes, storm water need to be treated. Biochar has been tried in storm water treatment and reuse by many investigators. The application of biochar supports the physical filtration and adsorption of pollutants. Table 5 indicates the removal efficiencies and adsorption capacities of pollutants from storm water onto biochars. In their study, Hasan et al. (2020) effectively utilized biochar-supported nanoscale zero-valent iron (BC-nZVI) for removing heavy metals (Cu, Cd, and Zn) from synthetic storm water at varying higher initial concentration ranges (2.5–60 mg/L) compared with typical urban storm water runoff. The maximum adsorption capacities of BC-nZVI for individual metal ions increased by 97% and 40% for Cd2+ and Zn2+, respectively, compared with original biochar. The mixture of sand and biochar reached the highest adsorption capacity (mg/g) of 223.62 Cd2+, 166.66 Zn2+, and 260.86 Cu2+ (Hasan et al., 2020). A few years ago, Lau et al. (2017) used surface-modified biochar from wood for removing E. coli from water and they achieved very good results. The original and modified biochar in their study removed 87.9%–99.8% of E. coli and the rate of E. coli remobilization varied from 0.04% to 4.44%. A study conducted by Shimabuku et al. (2016) used pelletized pine forestry waste to prepare biochar for abating sulfamethoxazole from surface water, storm water, and wastewater effluent. Their report indicated that the sulfamethoxazole sorption capacity of biochars produced by pyrolysis at 850°C from wood and wastewater sludge is comparable to a commercial powdered activated carbon. Xiong et al. (2019) revealed that iron-coated biochar effectively removed ammonia, phosphorus, and reduced COD parameters in water (Xiong et al., 2019).

5. Conclusions and perspectives Storm water has convincing potential to serve as an alternative source able to deal with the current shortage of clean water sources. Application of biochar could help remove pollutants such as heavy metals, organic pollutants, nutrients, antibiotics, and bacteria from storm water and make it available for irrigation or residential use. Biochar could be prepared from biomass via pyrolysis in a limited oxygen state and modified by many techniques and chemicals such as acids, bases, organic reagents, etc., depending on availability and target contaminants. Many studies have successfully developed biochars and applied them to storm water reuse scenarios. However, these are mostly preliminary results achieved in a laboratory or a small-scale setting. Thus, to apply storm water reuse on a larger scale, many tasks will have to be redesigned. Attempts should be made to produce more suitable biochar derived from waste biomass. More research in the large-scale application of biochar to reuse storm water is required. Furthermore, good collaboration

Table 5

Removal capacity of different filter media for storm water runoff treatment by Biochar.

Filter media

Description

Pollutant

Biochar

Carbon-dense solid residue from biomass pyrolysis Double surface area; wet sulfuric oxidation process Reduced porosity and enriched O-containing group density A mixture of sand and biochar

E. coli ATCC 10798 (synthetic storm water with E. coli concentration of 0.3–3.2  106 CFU/mL)

H2SO4-modified biochar Amino-modified biochar Biochar sand filter

Nanoscale zerovalent ironmodified biochar Biochar

Biochar

Biocharamended woodchip mixed woodchip Iron-coated biochar

A mixture of sand and biochar

E. coli ATCC 10798 (synthetic with a concentration of 1.2–1.7  106 CFU/mL) Heavy metal (Cu, Cd, and Zn)

Slow pyrolysis of pinewood biomass in an oxygen-deprived Produced around 850 °C

Heavy metal (Cu, Cd, and Zn)

Column experiments were employed to simulate a storm water treatment system Designed with doublelayer configurations

Heavy metal (Cd, Cu, Ni, Pb)

Sulfamethoxazole (100 ng/L to 1 mg/L)

COD, ammonia, phosphorus

Removal efficiency (%)

Mechanisms

References

96.6

Physical adsorption

Lau et al. (2017)

98.7

Pore straining and hydrophobic attraction

Lau et al. (2017)

92.1

Porosity reduction (mesopore)

Lau et al. (2017)

99.5–99.9

Stronger attachment/higher water-holding capacity

Mohanty et al. (2014)

Chemisorption process, chemical reduction, and surface complexation Chemisorption process, chemical reduction, and surface complexation

Hasan et al. (2020)

91.7% Cd2+; 94.2% Zn2+; >99% Cu2+ 48% Cd2+; 42% Zn2+

Qmax (mg/g)

223.62 Cd2+; 166.66 Zn2+; 260.86 Cu2+ 113.04 Cd2+; 118.76 Zn2+; 253.87 Cu2+

Around 35% at 60 min and 75% at 7 days >80%

94.6%; 98.3%; 93.70%

Hasan et al. (2020) Shimabuku et al. (2016)

Ashoori et al. (2019)

Iron filings were unable to adsorb phosphate; phosphate was mostly remobilized due to iron

Xiong et al. (2019)

Continued

Table 5

Removal capacity of different filter media for storm water runoff treatment by Biochar—cont’d

Filter media

Description

Pollutant

Rice husk biochar

Designed with doublelayer configurations Pyrolyzing granular poultry litter, which was made from raw poultry litter Made from Pennington Nature’s Heat hardwood pellets for household heating Pyrolyzed in a muffle at different temperatures under oxygen-limited condition

Ammonia, phosphorus

Poultry litter biochar

Hardwood biochar

Biochar wood dust for biocharamended biofilter with plant

Removal efficiency (%)

Qmax (mg/g)

57% 97%

Ammonium

1332.301 mg/kg

Ammonium

152.959 mg/kg

Bisphenol A, E. coli

>98.4%

Mechanisms

References

reduction, formation of metalsulfide precipitates Functional group of biochar (e.g., carboxylic and phenolic functional groups), increase its surface negative charge and decrease surface positive charge

Xiong et al. (2019) Tian et al. (2016)

Adsorption, filtration, plant uptake, and biodegradation

Lu and Chen (2018)

Tian et al. (2016)

Chapter 9 • Sustainable application of biochar for storm water reuse 273

between technological and policy drivers is the key to effective and widely accepted application of biochar in storm water reuse as a “green” solution which bring benefits to society, the economy and the environment. Recovery of fuels, nutrients and energy should be encouraged at the national level now and well into the future. Finally, commitment to making progress in national and international goals and awareness and active participation of all people should be enhanced.

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De Laurentiis, V., Corrado, S., Sala, S., 2018. Quantifying household waste of fresh fruit and vegetables in the EU. Waste Manag. 77, 238–251. Goonetilleke, A., Liu, A., Managi, S., Wilson, C., Gardner, T., Bandala, E.R., Walker, L., Holden, J., Wibowo, M.A., Suripin, S., Joshi, H., 2017. Stormwater reuse, a viable option: fact or fiction? Econ. Anal. Policy 56, 14–17. Han, T.N., Hoang, X.C., Pham, T.T.H., Nguyen, M.K., 2021. Study on acidity and neutralizing ability of ions in the chemical composition of rainwater. VNU J. Sci. Earth Environ. Sci. 37 (2), 98–106. Hasan, M.S., Geza, M., Vasquez, R., Chilkoor, G., Gadhamshetty, V., 2020. Enhanced heavy metal removal from synthetic stormwater using nanoscale zerovalent iron-modified biochar. Water Air Soil Pollut. 231 (5), 1–15. Hoang, M.T., Pham, T.D., Pham, T.T., Nguyen, M.K., Dang, T.T.N., Nguyen, T.H., Bartling, S., Bart, V.D.B., 2020. Esterification of sugarcane bagasse by citric acid for Pb2+ adsorption: effect of different chemical pretreatment methods. Environ. Sci. Pollut. Res. 28 (10), 11869–11881. Hong, N., Liu, A., Zhu, P., Zhao, X., Guan, Y., Yang, M., Wang, H., 2018. Modelling benzene series pollutants (BTEX) build-up loads on urban roads and their human health risks: implications for stormwater reuse safety. Ecotoxicol. Environ. Saf. 164, 234–242. Ilhan, Z.E., Ong, S.K., Moorman, T.B., 2012. Herbicide and antibiotic removal by woodchip denitrification filters: sorption processes. Water Air Soil Pollut. 223 (5), 2651–2662. Ioannidou, O.A., Zabaniotou, A.A., Stavropoulos, G.G., Islam, M.A., Albanis, T.A., 2010. Preparation of activated carbons from agricultural residues for pesticide adsorption. Chemosphere 80 (11), 1328–1336. Kalderis, D., Bethanis, S., Paraskeva, P., Diamadopoulos, E., 2008. Production of activated carbon from bagasse and rice husk by a single-stage chemical activation method at low retention times. Bioresour. Technol. 99 (15), 6809–6816. Kamala, C.T., Chu, K.H., Chary, N.S., Pandey, P.K., Ramesh, S.L., Sastry, A.R.K., Sekhar, K.C., 2005. Removal of arsenic (III) from aqueous solutions using fresh and immobilized plant biomass. Water Res. 39 (13), 2815–2826. Khoramzadeh, E., Nasernejad, B., Halladj, R., 2013. Mercury biosorption from aqueous solutions by sugarcane bagasse. J. Taiwan Inst. Chem. Eng. 44 (2), 266–269. Krishnani, K.K., Meng, X., Christodoulatos, C., Boddu, V.M., 2008. Biosorption mechanism of nine different heavy metals onto biomatrix from rice husk. J. Hazard. Mater. 153 (3), 1222–1234. Lau, A.Y., Tsang, D.C., Graham, N.J., Ok, Y.S., Yang, X., Li, X.D., 2017. Surface-modified biochar in a bioretention system for Escherichia coli removal from stormwater. Chemosphere 169, 89–98. Lee, J., Jeong, H., Choi, J.Y., Ra, K., 2020a. Characteristics and assessment of metal pollution and their potential source in stormwater runoff from Shihwa industrial complex, Korea. Korean J. Ecol. Environ. 53 (1), 91–101. Lee, J., Jeong, H., Ra, K., Choi, J.Y., 2020b. Assessment of particle size distribution and pollution impact of heavy metalsin road-deposited sediments (RDS) from Shihwa industrial complex. J. Environ. Impact Assess. 29 (1), 8–25. Li, H., Wang, Y., Liu, F., Tong, L., Li, K., Yang, H., Zhang, L., 2018. Volatile organic compounds in stormwater from a community of Beijing, China. Environ. Pollut. 239, 554–561. Liu, P., Liu, W.J., Jiang, H., Chen, J.J., Li, W.W., Yu, H.Q., 2012. Modification of bio-char derived from fast pyrolysis of biomass and its application in removal of tetracycline from aqueous solution. Bioresour. Technol. 121, 235–240. Liu, A., Gunawardana, C., Gunawardena, J., Egodawatta, P., Ayoko, G.A., Goonetilleke, A., 2016. Taxonomy of factors which influence heavy metal build-up on urban road surfaces. J. Hazard. Mater. 310, 20–29. Liu, A., Ma, Y., Gunawardena, J.M., Egodawatta, P., Ayoko, G.A., Goonetilleke, A., 2018. Heavy metals transport pathways: the importance of atmospheric pollution contributing to stormwater pollution. Ecotoxicol. Environ. Saf. 164, 696–703.

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Lu, L., Chen, B., 2018. Enhanced bisphenol A removal from stormwater in biochar-amended biofilters: combined with batch sorption and fixed-bed column studies. Environ. Pollut. 243, 1539–1549. Luthy, R.G., Sharvelle, S., Dillon, P., 2019. Urban stormwater to enhance water supply. Environ. Sci. Technol. 53, 5534–5542. Mahbub, P., Goonetilleke, A., Ayoko, G.A., 2011a. Prediction model of the buildup of volatile organic compounds on urban roads. Environ. Sci. Technol. 45 (10), 4453–4459. Mahbub, P., Goonetilleke, A., Ayoko, G.A., Egodawatta, P., 2011b. Effects of climate change on the wash-off of volatile organic compounds from urban roads. Sci. Total Environ. 409 (19), 3934–3942. Milik, J., Pasela, R., 2018. Analysis of concentration trends and origins of heavy metal loads in stormwater runoff in selected cities: a review. In: E3S Web of Conferences. vol. 44. EDP Sciences, p. 00111. Mohanty, S.K., Cantrell, K.B., Nelson, K.L., Boehm, A.B., 2014. Efficacy of biochar to remove Escherichia coli from stormwater under steady and intermittent flow. Water Res. 61, 288–296. Mohanty, S.K., Valenca, R., Berger, A.W., Iris, K.M., Xiong, X., Saunders, T.M., Tsang, D.C., 2018. Plenty of room for carbon on the ground: potential applications of biochar for stormwater treatment. Sci. Total Environ. 625, 1644–1658. Moreno, R.F., Lo´pez, C.J., Galvı´n, M.R., Cordo´n, M.M., Mellado, R.J., 2010. On the removal of s-triazine herbicides from waters using commercial low-cost granular carbons. J. Serb. Chem. Soc. 75 (3), 405–412. Mullaugh, K.M., Hamilton, J.M., Avery, G.B., Felix, J.D., Mead, R.N., Willey, J.D., Kieber, R.J., 2015. Temporal and spatial variability of trace volatile organic compounds in rainwater. Chemosphere 134, 203–209. Nasernejad, B., Zadeh, T.E., Pour, B.B., Bygi, M.E., Zamani, A., 2005. Comparison for biosorption modeling of heavy metals (Cr (III), Cu (II), Zn (II)) adsorption from wastewater by carrot residues. Process Biochem. 40 (3–4), 1319–1322. Nguyen, M.K., Hoang, M.T., Pham, T.T., Bart, V.D.B., 2019. Performance comparison of chemically modified sugarcane bagasse for removing Cd (II) in water environment. J. Renew. Mater. 7 (5), 414–427. ` ., Jove
, P., Oliveras, A., 2011. The use of cork waste as a biosorbent for persistent organic Olivella, M.A pollutants—study of adsorption/desorption of polycyclic aromatic hydrocarbons. J. Environ. Sci. Health A 46 (8), 824–832. Park, K.H., Balathanigaimani, M.S., Shim, W.G., Lee, J.W., Moon, H., 2010. Adsorption characteristics of phenol on novel corn grain-based activated carbons. Microporous Mesoporous Mater. 127 (1–2), 1–8. Payne, E.G., McCarthy, D.T., Deletic, A., Zhang, K., 2019. Biotreatment technologies for stormwater harvesting: critical perspectives. Curr. Opin. Biotechnol. 57, 191–196. Pino, G.H., de Mesquita, L.S., Torem, M.L., Pinto, G.A.S., 2006. Biosorption of heavy metals by powder of green coconut shell. Sep. Sci. Technol. 41 (14), 3141–3153. Rahman, M.Y.A., Nachabe, M.H., Ergas, S.J., 2020. Biochar amendment of stormwater bioretention systems for nitrogen and Escherichia coli removal: effect of hydraulic loading rates and antecedent dry periods. Bioresour. Technol. 310, 123428. Ranjan, D., Talat, M., Hasan, S.H., 2009. Rice polish: an alternative to conventional adsorbents for treating arsenic bearing water by up-flow column method. Ind. Eng. Chem. Res. 48 (23), 10180–10185. Shimabuku, K.K., Kearns, J.P., Martinez, J.E., Mahoney, R.B., Moreno-Vasquez, L., Summers, R.S., 2016. Biochar sorbents for sulfamethoxazole removal from surface water, stormwater, and wastewater effluent. Water Res. 96, 236–245. Strauch, A.M., MacKenzie, R.A., Giardina, C.P., Bruland, G.L., 2015. Climate driven changes to rainfall and streamflow patterns in a model tropical island hydrological system. J. Hydrol. 523, 160–169. Sun, Y., Chen, S.S., Lau, A.Y., Tsang, D.C., Mohanty, S.K., Bhatnagar, A., Rinklebe, J., Lin, K.Y.A., Ok, Y.S., 2020. Waste-derived compost and biochar amendments for stormwater treatment in bioretention column: co-transport of metals and colloids. J. Hazard. Mater. 383, 121243.

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Tian, J., Miller, V., Chiu, P.C., Maresca, J.A., Guo, M., Imhoff, P.T., 2016. Nutrient release and ammonium sorption by poultry litter and wood biochars in stormwater treatment. Sci. Total Environ. 553, 596–606.
rez, J., Ge
rente, C., Andre`s, Y., 2012. Sustainable activated carbons from agricultural residues Torres-Pe dedicated to antibiotic removal by adsorption. Chin. J. Chem. Eng. 20 (3), 524–529. Tran, V.S., Ngo, H.H., Guo, W., Zhang, J., Liang, S., Ton-That, C., Zhang, X., 2015. Typical low cost biosorbents for adsorptive removal of specific organic pollutants from water. Bioresour. Technol. 182, 353–363. Tran, V.C., Pham, Q.H., Vu, D.T., Nguyen, M.K., 2020. Influence of biochar ammendments on surface charge and bioavailability of heavy metals in degraded soils. Suranaree J. Sci. Technol. 27 (4), 020008-3. Wang, L.K., Hung, Y.T., Shammas, N.K. (Eds.), 2007. Advanced Physicochemical Treatment Technologies. vol. 5. Humana Press, Totowa, NJ, pp. 575–610. Weerasundara, L., Magana-Arachchi, D.N., Ziyath, A.M., Goonetilleke, A., Vithanage, M., 2018. Health risk assessment of heavy metals in atmospheric deposition in a congested city environment in a developing country: Kandy City, Sri Lanka. J. Environ. Manage. 220, 198–206. https://www-sciencedirect-com. db.lic.vnu.edu.vn/science/article/pii/S0301479718304183. Wijesiri, B., Liu, A., Egodawatta, P., McGree, J., Goonetilleke, A., 2019. Understanding Uncertainty Associated with Stormwater Quality Modelling. In: Decision Making with Uncertainty in Stormwater Pollutant Processes. Springer, Singapore, pp. 1–13. https://link.springer.com/chapter/10.1007/978981-13-3507-5_1. Wong, K.K., Lee, C.K., Low, K.S., Haron, M.J., 2003. Removal of Cu and Pb by tartaric acid modified rice husk from aqueous solutions. Chemosphere 50 (1), 23–28. Xi, Z., Chen, B., 2014. Removal of polycyclic aromatic hydrocarbons from aqueous solution by raw and modified plant residue materials as biosorbents. J. Environ. Sci. 26 (4), 737–748. Xiong, J., Ren, S., He, Y., Wang, X.C., Bai, X., Wang, J., Dzakpasu, M., 2019. Bioretention cell incorporating Fe-biochar and saturated zones for enhanced stormwater runoff treatment. Chemosphere 237, 124424. Yaduvanshi, A., Ranade, A., 2015. Effect of global temperature changes on rainfall fluctuations over river basins across eastern Indo-Gangeticplains. Aquat. Procedia 4, 721–729. Yakout, S.M., Daifullah, A.A.M., 2013. Removal of selected polycyclic aromatic hydrocarbons from aqueous solution onto various adsorbent materials. Desalin. Water Treat. 51 (34–36), 6711–6718. Yu, S., Lee, P.K., Yun, S.T., Hwang, S.I., Chae, G., 2017a. Comparison of volatile organic compounds in stormwater and groundwater in Seoul metropolitan city, South Korea. Environ. Earth Sci. 76 (9), 1–17. Yu, Y., Wang, W., Shi, J., Zhu, S., Yan, Y., 2017b. Enhanced levofloxacin removal from water using zirconium (IV) loaded corn bracts. Environ. Sci. Pollut. Res. Int. 24 (11), 10685. Zolgharnein, J., Shahmoradi, A., Ghasemi, J., 2011. Pesticides removal using conventional and low-cost adsorbents: a review. Clean Soil Air Water 39 (12), 1105–1119.

Further reading Ata, A., Nalcaci, O.O., Ovez, B., 2012. Macro algae Gracilaria verrucosa as a biosorbent: a study of sorption mechanisms. Algal Res. 1 (2), 194–204. Mahanta, R., Chowdhury, J., Nath, H.K., 2016. Health costs of arsenic contamination of drinking water in Assam, India. Economic Analysis and Policy 49, 30–42.

10 Biochar for sustainable remediation of soil Yuanyao Yea, Huu Hao Ngob, Wenshan Guob, Jianxiong Kanga, Wei Jianga, Yongzheng Rena, and Dongqi Liua a SCHOOL OF ENVIRONMENTAL SCIENCE AND E NGI NEERI NG, HUAZHONG UNIVERSITY O F SCIENCE AND TECHNOLOGY, WUHAN, PR CHINA b CENTRE FOR T ECHNOLOGY IN WATER AND WASTEWATER, S CHOO L O F CIVIL AND ENVIRONMENTAL E NGINEERING, UNIVERSITY O F TECHNOLOGY SYDNEY, SYDNEY, NS W, AUSTRALIA

1. Introduction Currently, soil plays a crucial role in human survival and development since 90% of human sources including fuel, fiber, livestock, and food are highly dependent on soil. More specifically, soil can act as a habitat and water purifier for billions of living organisms, and provide groundwater and raw materials. For the 17 Sustainable Development (UNSDP) Goals proposed by United Nations, the soil has great impacts on realizing many of those goals, particularly for good health and wellbeing, improvement in sanitation, clean water, and food security. However, the explosive increase in human population and rapid growth in industrialization and urbanization have resulted in soil pollution over the last few decades (Ali et al., 2019; Tang et al., 2014). Typical soil pollutants include chlorophenols, herbicides, pesticides, heavy metal/metalloids (HMs), crude oil and its petrochemical derivatives, and polycyclic aromatic hydrocarbons (PAHs) (Chen et al., 2015). Hence, it is important to look for sustainably low-cost, environmentally friendly and effective remediation strategies to efficiently remediate and control the contaminants in soils. Even though existing chemical technologies (e.g., soil washing, electro-kinetics, immobilization, ion-exchange and precipitation) and conventional physical processes (e.g., soil replacement and thermal treatment) can remove the pollutants mentioned above from soil to some extent (Qayyum et al., 2020), the application of these methods is still subjected to poor feasibility, high cost, inefficiency, unsustainability and high secondary risk (Khalid et al., 2017; Lahori et al., 2017). Besides, some technologies may be only effective under certain conditions (Sabatini and Knox, 1992), result in loss of valuable metals and compromise soil quality (Hou and Gao, 2003; Khalid et al., 2017). For example, encapsulation and stabilization and solidification (S/S) cannot remove, transform and destroy contaminants even though those contaminants can be entrapped and stabilized in a solid

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form (Khalid et al., 2017). In contrast, bioremediation has attracted increased interest in recent years due to its being more environmentally friendly and economic in soil management. However, this technique’s reliability is challenged and it involves long treatment time. Furthermore, some exotic species may be introduced to remove the certain pollutants in the bioremediation process, which may pose risks to native biodiversity (Chibuike and Obiora, 2014). In recent times, biochar remediation has been considered a relatively cost-effective and environmentally beneficial tool for directing toward soil pollution. The possible reason for this is that the properties of biochar including negative surface charge density, high surface, resistance to degradation and porous structure facilitate nutrient affinity, water retention and soil remediation (Metcalf, 2016; Rajapaksha et al., 2016). Biochar is a carbon-containing organic solid material with a high degree of aromatization and strong antidecomposition properties (Pandey et al., 2020). Due to biochar’s properties such as porosity and abundant surface functional groups, the adsorption of organic and inorganic contaminants can be improved (Zhu et al., 2018). The removal of organic and inorganic pollutants from the soil through biochar is achieved by the adsorption and interactions between active functional groups on the biochar surface and these pollutants (Kamali et al., 2020; Mohamed et al., 2016; Tan et al., 2015). Moreover, the oxygen-containing functional groups, higher cation exchange capacity (CEC) and larger surface area of biochar can stabilize different heavy metal ions in soil and the bioavailability of these heavy metals to terrestrial floras can be reduced as well as their resultant toxicity (Beesley et al., 2011; Park et al., 2011). The removal behavior of biochar differs from contaminants and well correlated with the properties of contaminants (Guo et al., 2020b). Consequently, it is highly recommended to use biochar for soil amendment, especially because of its high efficiency, cost-effectiveness and ecofriendly nature for contaminant remediation (Dawood et al., 2017). Biochar is achieved through hydrothermal and thermochemical processes (e.g., pyrolysis and gasification), in which such organic materials can be converted into porous carbonaceous materials (Aziz et al., 2020; Sajjad et al., 2020). According to the life cycle assessment of pyrolysis-biochar systems, biochar prepared from waste biomass is more cost-effective and comparatively eco-friendly (Cowie et al., 2012). Besides, various publications have confirmed that engineered biochar can be successfully applied to energy storage, waste recycling, as a catalyst for biofuel production, environmental remediation and carbon sequestration (Tan et al., 2017). Its application in the soil can not only capture and then sequestrate greenhouse gases such as carbon dioxide to mitigate global warming, but also enhance soil properties including reduction of leaching soluble nutrients, and improvement in soil fertility, water holding capacity, aeration and microbial activity through modifying nutrient in the soil and microbial life (Giagnoni et al., 2019; Koga et al., 2017; Kumar et al., 2020; Liu et al., 2018; Wang et al., 2019; Zeeshan et al., 2020). As a result of this, the promising amendment is beneficial for remediating contaminated soils and the plant growth and crop yield are positively affected. The generation of biomass waste such as food waste and forestry/agricultural wastes rapidly increases over the last decade, which is attributed to the rapid growth in living

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standards, global population and urbanization (Salehi et al., 2020; Yang et al., 2020). Therefore, the conversion of biomass to biochar is not only a promising strategy for soil remediation, but also an environmentally acceptable solution for sustainable biomass waste management and recycle of forestry and agricultural wastes (El-Bassi et al., 2021;
´k, 2021; Salehi and Hosseinifard, Gujre et al., 2020; Jung et al., 2019; Kwoczynski and Cmelı 2020; Singh et al., 2020). Various raw materials have been used for biochar preparation, including crop residues, woodchips, animal manure, lignocellulosic plant materials, sawdust, rice straw, pine-wood, safflower seeds, wood and bamboos (Angın, 2013; Fan et al., 2010; Qian et al., 2017; Senthil and Lee, 2020; Sun et al., 2020; Vaughn et al., 2013; Wang et al., 2015; Zhou et al., 2017). Table 1 presents the application of biochar for soil

Table 1 Biomass feedstock

Soil remediation by biochar application. Soil texture

Remediation performance

Reference

Grape prune residue Sludgebased biochar

Calcisols

The mobility factor of Zn, Cu, Pb, and Cd reduced by 49%, 70%, 62%, and 47%, respectively

Sandy soil and loamy soil

Corn straw

Sandy loam soil

Peanut vine and rice straw Corn stalk

Latosolic red soil

Wheat straw

Topsoil without impurities such as gravel and plant roots

Spartina alterniflora

The soil composed of perlite, coconut bran, kaolin, and turfy earth

The content of Cu, Zn, As, Pb, Cd, and Cr decreased to 9.44%, 63.34%, 52.11%, 4.87%, 12.5%, and 9.57%, respectively The activities of soil microbes in loamy soil and sandy soil, root length and the shoot length can be increased by 154.74%–195.76%, 92.31%– 157.69%, 66.81%–96.45% and 37.5%–53.32%, respectively 93.2% of benzo(a)pyrene can be removed from the soil The content of exchangeable Cd content in the two biochars was decreased by 35.80% and 28.48%, respectively The soil pH increased by 0.31 The content of Pb, Zn, Cu, and Cd decreased by 23.8%, 5.27%, 14.3%, and 11.9%, respectively The activities of antioxidant enzymes were inhibited by 29.6%–51.3% The biodegradation efficiency of benzo(a)pyrene was 35.4%–86.6% after adding the biochar The activity of polyphenol oxidase and dehydrogenase increased by 26.6%–69.1% and 27.5%–70.2%, respectively The bioconcentration factor values in B. chinensis were 24%–31% The recalcitrance of the soil amendment was improved while the Cd bioavailability was increased

Taghlidabad and Sepehr (2017) Xing et al. (2019)

Vegetable field

Guo et al. (2020a) Chen et al. (2020) Huang et al. (2020)

Guo et al. (2021)

Qiu et al. (2020)

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Food waste

Forestry/Agricultural wastes

Manure

Biomass

Pyrolysis

Residues

Soil

Biochar

Remediation FIG. 1 Relationship among biomass, biochar and soil remediation.

remediation while the relationship among biomass, biochar and soil remediation is depicted in Fig. 1.

2. Potential effects of biochar on soil Biochar can be effectively and efficiently utilized by microorganisms in soil due to containing the easily degradable carbon that can be washed and mineralized in the soil (Roberts et al., 2015). Thus, additional biochar in the soil can increase the soil enzyme activity, the amount of microbial biomass and soil biomass, which facilitates the remediation of soil (Abbas et al., 2018). Besides, biochar has high economic and ecological benefits, including the increase in crop yields and improvement in soil quality (Ouyang et al., 2016; Plaza et al., 2016). The effects of biochar on soil are presented in Fig. 2.

2.1 Biological properties Living organisms in the soil can be acted as an indicator to reflect the changes in ecosystem functions and thus provide a reliable basis for soil properties. Besides, the biochar addition to the soil may influence the physicochemical properties of soil as well as the microbial activity in soil (Sun et al., 2018). Biochar’s surface properties and highly aromatic hydrocarbon structure easily make itself become a habitat for soil microbial organisms such as fungi, bacteria and algae (Knicker, 2007).

2.1.1 bacteria Bacterial decomposition can easily happen due to the existence of easily decomposed N and C sources on the biochar surface. As a result of this, bacteria can be adsorbed to

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FIG. 2 The effects of additional biochar on soil.

the biochar surface and the number of bacteria in soils is increased. This is because bacteria are less susceptible to soil leaching after adding biochar (Pietikainen et al., 2000). Kim et al. (2007) reported the application of biochar could increase the diversity by 25% and enhance the nitrogen fixation capacity and the number of nitrogen fixing bacteria. Besides, the biochar addition can also promote bacterial nitrification (Gundale and Deluca, 2006) and increase the relative abundance of bacteria, but decrease the soil microbial diversity (Khodadad et al., 2011). Purakayastha et al. (2019) also found that the application of sewage sludge biochar to loamy sand can increase the percentage of actinomycetes, gram-negative bacteria and gram-positive bacteria. Appropriate addition of biochar can enhance the activity of nitrogen-fixing bacteria, which facilitates the conversion of N2 in the atmosphere into nitrates (NO3 ) via nitrification. As the application rate of biochar to soil increased from 30 and 60 g/kg, the N content in soil increased by 29% (Rondon et al., 2007). However, the N content was reduced to 30% while continuously adding the biochar (60 g/kg). It indicates that the impacts of biochar on soil bacteria are complicated, which may be determined by the additive amount of biochar, preparation condition of biochar and other possible factors.

2.1.2 Fungus The addition of biochar to soil could not only provide a habitat for soil microbial growth, but also protect soil microorganisms and reduce the competition among microorganisms (Warnock et al., 2010). The application of biochar could make significant improvement in the fungal reproduction when compared to the bacterial diversity (Luo et al., 2017). The biochar addition could increase the biomass of fungi in the soil (Antonietti, 2009), but the excessive application of biochar may have negative impacts on the amount of fungus (Nie

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et al., 2018). Besides, biochar in soil increases the supply of water and nutrient, and reduces the demand for mycorrhizal symbiosis, for which the abundance of fungi can be reduced while the effectiveness of P can be increased after adding biochar (Warnock et al., 2010). Since the biochar production varied from thermal cracking conditions, technical processes and raw materials, the effects of biochar on soil ecosystem functions should be evaluated under specific conditions.

2.1.3 Microbial biomass The application of biochar could significantly change the microbial structure. The possible reason for this is that partially soluble N and C sources on the biochar surface are beneficial to microbial activity while the large specific area and pore structure of biochar can store nutrients and water, which forms a microenvironment for organisms to live (Ema et al., 2019). Furthermore, the circulation of soil organic matter and biome are influenced by the biochar addition (Clough et al., 2010). Ema et al. (2019) noted that the application of biochar to soil can increase the organic matter in agricultural soils as well as the number of NO3 -N. Moreover, the addition of biochar may increase the amount of soil microorganisms (Wardle et al., 2008), but some studies have shown that the impacts of biochar on the soil microbial biomass may rely on the type of biochar (Castaldi et al., 2011; Dempster et al., 2012). Besides, the composition and number of microbial communities may be harmed by the high application rates of biochar (You et al., 2019). The impacts of biochar on soil microbes also depend on the nature of biochar, nutrient management, land use patterns, fertility status and soil texture.

2.1.4 Soil enzyme activity Soil enzyme activity can reflect the direction and intensity of various biochemical processes in soils (Awad et al., 2012; Zhang et al., 2021). The impacts of biochar addition on agricultural soil microecology can be analyzed by the changes in the soil metabolic properties (Zhang et al., 2021). Pandey et al. (2016) found that urease activity increased from 7.4% to 39%. Besides, compared with the control group, the biochar addition resulted in an increase of soil enzyme activity (dehydrogenase) by 27%. Inconsistent activity changes in acetylglucosamine, leucine aminopeptidase, lipase, and glucoamylase in different soils may happen, which is attributed to the impacts of biochar addition on the adsorptioninhibiting enzyme reaction in soils (Bailey et al., 2011). It was reported that there is a specific relationship between the amount of biochar added and enzyme activity (Awad et al., 2012). Less demand for the enzyme production occurs due to the increased colocalization efficiency of carbon on the biochar surface and microbes (Smith et al., 2010).

2.2 Chemical properties 2.2.1 Organic matter Biochar addition could increase the efficiency of plants in absorbing nutrients, the effectiveness of soil nutrients and the soil organic matter content, in which the amount and

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stability of biochar highly affect its effects (Oz, 2018; Jeffery et al., 2011). The content of mineral nutrients such as K, N, Mg, and P and some cations such as Cu, Zn, and Ca in soils can be increased after the addition of biochar, which improves the availability of soil nutrients and thus facilitates the plant growth (Smith et al., 2010). It should be noted here that higher soil organic matter contributes to the crop yields since the absorption of organic matter by roots can be improved by adding biochar as well as the activity of root microorganisms (Liang et al., 2014). This is attributed to the increase in the soil bulk density (SBD) reduction, available soil nutrients and soil porosity (Akhtar et al., 2014). Some field trials have also confirmed that the biochar application could increase the yields of crops and plants (Major et al., 2010; Steiner et al., 2007). In addition to the crop yield growth, the biochar addition could effectively prevent the plant from being infected by pests and enhance plant disease resistance (Kumar et al., 2018). More importantly, the biochar addition may have detrimental impacts on the crop yields due to its inhibition effects on the original productivity of the soil. Thus, the biochar effects on the crop yields depend on the soil’s nature and the amount of biochar added.

2.2.2 Nitrogen The conversion, retention and circulation of N can be affected by the biochar addition, where biochar can effectively adsorb ammonium nitrogen (Taghizadeh-Toosi et al., 2012). As a result of this, more nitrogen can be reused by plants after the biochar addition, € eren˜a et al., 2013). thus reducing the leaching of soil N and increasing the N availability (Gu However, Yi et al. (2012) found that biochar application may negligibly influence the leaching of nitrate nitrogen. In sand culture experiments, biochar had the best adsorption performance for ammonium nitrogen at neutral pH, but this effect is highly affected by the pH value (Chen et al., 2013). Asada et al. (2002) found that the number of acidic functional groups on the biochar surface is decreased at higher pyrolysis temperature during the biochar preparation, in which the NH3 adsorption on biochar is seriously affected. Moreover, incomplete denitrification can be improved by adding biochar which is generated at low temperature (Weldon et al., 2019). To sum up, the impacts of biochar addition on the adsorption of soil NH4+ and NH3 are governed by the pH of the applied soil, the additive amount and the preparation temperature of the biochar.

2.2.3 Carbon Adding biochar in soil facilitates the carbon capture and further sequester in the soil, lowering the carbon emission. More specifically, the biochar application in the soil can inhibit the methane emission (Liu et al., 2011), where the CH4 emissions can be reduced by 96% (Karhu et al., 2011). In contrast, a lab-scale field trial indicated that the addition of biochar has negligible impacts on the inhibition of CH4 emission while adding 16 different kinds of biochars to three types of soils (Spokas and Reicosky, 2009). This may be attributed to the difference in the biochar properties and soil nature. Moreover, Alho et al. (2012) found that low addition of biochar may enhance the N2O release, but increasing the amount of biochar added to over 5 mg/ha could mitigate the N2O emission.

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2.2.4 CEC The CEC is employed to assess the ability of soil to exchange, retain and adsorb cations, in which the soil CEM content can be increased by increasing the number of soil cation exchange sites. High soil CEM content enables the soil to more efficiently adsorb Mg2+, Ca2+, K+, and NH+4 , which facilitates the retention and utilization of nutrient ions in the soil. Studies show that the biochar addition can result in increasing the content of CEC and soil charge by about 20%–40% compared to the control group (Liang et al., 2006). This is despite the fact that the amount of nutrients and alkaline cations can be significantly increased by adding a small amount of biochar (Hossain et al., 2010). However, the biochar addition may not have any effects on the CEC of soils containing high organic content since such soils already have high CEC content (Schulz, 2012). Furthermore, both the pyrolysis conditions and raw biomass feedstocks can affect the CEC of produced biochar. For example, increasing the temperature from 450°C to 700°C could result in the reduction in the CEC from 26.36 to 10.28 cmol/kg (Lee et al., 2010).

2.2.5 pH The pH of biochar ranges from 4 to 12 and its alkaline property has direct impacts on soil pH after its addition. The soil pH value can be regulated by the biochar addition via increasing the saturation of the base (Hossain et al., 2010; Nielsen et al., 2018). As the amount of biochar added increases, the pH may gradually increase, but no effects on the pH of alkaline soils are observed after the biochar application (Chintala et al., 2014). Under the water action, biochar can exchange with Al3+ and H+ in the soil to decrease those ions’ concentration (Zwieten et al., 2010).

2.3 Physical properties 2.3.1 Soil bulk density SBD is an important indicator of soil physical properties, which is closely related to the tightness of soil. Laird et al. (2010) reported that the application of biochar in the soil can significantly decrease SBD, which can effectively reduce soil compaction, facilitate nutrient release and retention, and improve soil structure (Chen et al., 2010). In this scenario, the SBD of silty soil was reduced from 1.52 to 1.33 g/cm3 after adding biochar. Studies have shown that the biochar addition can increase total porosity, enhance soil agglomeration, and promote soil fungal growth (Qin et al., 2016; Steiner et al., 2007).

2.3.2 Soil porosity The biochar’s particle size, connectivity and pore distribution have great impacts on the soil pore structure, where the soil pore in turn significantly influences the utilization, storage and transformation of water, and offers the necessary oxygen and space for fauna and soil flora. The application of biochar containing large surface area contributes to plant root growth and microbial activities in soil pores and results in increasing the soil microorganisms. Oguntunde et al. (2010) found that the biochar addition can lead to the growth

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of the soil saturated hydraulic, soil permeability and soil porosity. The crop productivity increases by 13% and 10% after adding the biochar to the soils with a medium or coarse texture, respectively ( Jeffery et al., 2011). The possible reason for this is that adding biochar can facilitate the reduction in the soil pore size and have positive effects on crop growth (Dokoohaki et al., 2017). Rasa et al. (2018) found that the soil porosity ranges from 5 to 10–25 μm after adding biochar. In summary, the biochar addition can improve soil fertility, water conservation and soil compaction, increase air and water circulation, improve soil pore connectivity and change pore distribution and soil porosity.

2.3.3 Soil hydraulic properties The biochar application in the soil can change the water residence time and flow path, soil water holding capacity, and reduce soil water permeability resistance due to its e high porosity and large specific surface area (Abrol et al., 2016; Major et al., 2012). Studies have shown that the biochar addition can increase the soil water holding capacity in the field and thus improve the crop yield (Verheijen et al., 2019; Yang et al., 2015). Moreover, the use of biochar can improve the water use efficiency of plants, so it is recommended to employ biochar in areas with scarce water sources (Fischer et al., 2019). The porosity, connectivity, shape, distribution and pore size of soil can be changed due to the biochar application, for which the mechanical strength and water retention of soil can be affected. Biochar with average pore diameter and larger pore volume may have better water retention. This indicates that the biochar addition can enhance soil water hydraulic conductivity. The saturated hydraulic conductivity can be increased after adding biochar to soil (MoraguesSaitua et al., 2017; Trifunovic et al., 2018). Soil can accelerate drainage and infiltration when the soil has high water conductivity (Abel et al., 2013). Even though the rapid drainage of the soil can effectively decrease the risks of runoff, it may reduce the amount of agrochemicals and nutrients available for being filtered and dissolved in water (Li et al., 2013). The impacts of biochar on soil saturated hydraulic conductivity are different, which is attributed to the difference in the amount of biochar added and the different types of biochar’s surface and pore structures (Uzoma et al., 2011; Zhang et al., 2016).

2.3.4 Soil temperature and color Biochar addition can influence the soil temperature and surface reflectance, for which the soil heat can be affected (Briggs et al., 2012). It was reported that adding biochar to the soil can increase the soil temperature (Ventura et al., 2012; Zhang et al., 2013). Besides, the biochar application in the soil can also darken the soil color due to its being a black particulate matter.

2.4 Environmental risks The biochar application as soil amendment may also pose risks on environmental and ecological systems (Doaa et al., 2021; Gwenzi et al., 2015). Some raw biochar materials

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such as polluted crops and livestock manure may contain heavy metals that are difficult to be removed during the biochar production via pyrolysis (Schievano et al., 2016). Consequently, the heavy metals may retain in carbonized products and thus lead to pollution and negatively influence the adsorption capacity of biochar (Uchimiya et al., 2010; Zheng et al., 2012). Biochar aging can detrimentally affect the physicochemical properties of biochar (Peng et al., 2020), increase the heavy metal mobility and release of dissolved organic matter (Kim et al., 2021). Even though the positive results of biochar application in soil have been widely reported, some studies indicate that the application of some biochar may have serious effects on microbial activity and soil remediation (El-Naggar et al., 2019). An overview of the possible interactions between the additional biochar and soil is shown in Fig. 3.

FIG. 3 An overview of the possible interactions between the additional biochar and soil (Biswal et al., 2022; Zhang et al., 2018).

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3. Biochar and bioeconomy Bioeconomy is to use biotechnology to covert waste bioresources into economically valuable bioproducts through biotechnology (Bugge et al., 2016; Stegmann et al., 2020). Some developed countries have adopted the bioeconomy to refine and upgrade bioresources (Bugge et al., 2016; Maestrini et al., 2014). Nevertheless, it is essential to encourage bioproduct production and raise the awareness of bioeconomy. In this scenario, biochar produced from various biosources can add value to the bioresource promptly and be used for environmental and agricultural applications (Budzianowskiab, 2017; Leng et al., 2012). Compared to the traditional take-make-dispose linear model, the concept of bioeconomy in principle promotes sustainability. Furthermore, the bioeconomy is highly affected by the quantity, safety and quality of bioproduct (i.e., biochar) (Fuentes-Saguar et al., 2017), which demonstrates that the biochar application, soil type and crop choice are very important due to their effects on the economic balance (Baidoo et al., 2016; Shareef and Zhao, 2017). Notably, the biochar application can not only be conducted in the energy, industry and agriculture sector, but also increase employment opportunities. Adding biochar to cultivate cash crops in agriculture can yield high returns. This is because the biochar application can improve the soil fertility and thus enhance the crop yield. Compared to the commercial fertilizers, there is no need to add biochar annually due to its stable persistence. Simultaneously, the use of biochar can improve the retention of water and soil, which recovers depleted soils, and lowers costs of fertilizers and irrigation. The utilization of biochar for soil remediation has been proved to be effective both conceptually and at short-term studies at lab-scale, which depends on the biochar dosage, pyrolysis conditions and type of raw feedstock materials (Bian et al., 2013; Chen et al., 2019; Hassan et al., 2020). At the lab-scale research, the biochar application for soil remediation has been proved to be sustainable and economically feasible (Gwenzi et al., 2017; Xiang et al., 2020) because (1) no transportation costs are involved in the small-scale study; (2) the biochar can be regenerated and reused; and (3) the biochar is derived from the free or cheap renewable waste biomasses. Apart from this, by controlling the specific pyrolysis conditions, type of carbonization methods, aging conditions, posttreatment measures and type of biochar, it may avoid or mitigate the secondary environmental risk related to the biochar application, which enhances the economic feasibility of the soil remediation system (Buss et al., 2015; Oleszczuk and Kołtowski, 2018). In addition, the economic assumption of the biochar production was US$ 532.00/year in Selangor, and the total revenue from biochar sale was US$ 8012/year according to the study of Harsono et al. (2013). Therefore, the net revenue and amount of investment determine the net present value for biochar production, demonstrating high economic feasibility of biochar. Furthermore, the selling price directly affects the cost-effectiveness of biochar production, where the biochar prepared at 450°C and 300°C had the break-even of around $280/t and $220/t, respectively (Shabangu et al., 2014). When the yard waste is used to produce biochar, the net margin is US$ 16 and 69 for the low and high revenue scenario of CO2e (Roberts et al., 2010). Transportation is another important factor to

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evaluate the economic feasibility of biochar preparation and application (Palma et al., 2011). Therefore, biochar generation would be attractive if its economic benefits can offset the economic costs of biochar’s application and transportation, pyrolysis process, and storing, hauling and harvesting of the raw biomass feedstocks. More importantly, the development in biochar preparation technology can improve the net margin of biochar application as well as the selection of more economical feedstock (Bugge et al., 2016).

4. Conclusions and perspectives Biochar which is obtained from various raw biomass materials can be used to improve the removal efficiency of pollutants in soils to satisfy soil remediation. Biochar is pyrolyzed at different conditions such as temperature, presenting highly heterogeneous chemical and physical characteristics. The type of feedstock and pyrolysis condition determines the quality and quantity of biochar. Besides, the biochar application has many positive impacts on soils, which enhances the soil remediation and crop yield. This is despite the fact that some pollutants may be retained in biochar and thus detrimentally influence the soils and plants. The use of biochar is an important part of bioeconomy because it improves the sustainable waste management, provides more opportunities for additional income, and helps soil remediation. Some knowledge gaps in biochar technology are still yet to be developed. Even though biochar technology for sustainable soil remediation has been widely studied at laboratory and small-scale field trials, the sustainability, cost, effectiveness and practicality still challenge its large-scale field application. It is essential to implement large-scale field trials before performing full-scale remediation projects. Apart from this, the costs involving equipment purchases, transportation (raw biowaste and biochar) and production (raw biowaste and pyrolysis) are another hurdle for conducting fullscale applications. Simultaneously, the long-term addition of biochar should conduct risk assessment to improve plant yield with the minimum risks. To improve the economic feasibility of the biochar application, it is recommended to produce biochar near the feedstock area to decrease the logistical costs. Furthermore, it is important to look for those free or cheap readily and locally available waste biomasses which can prepare biochar with reusability cycles and good regeneration performance. The integration of biochar and other degradation materials/organisms can improve the efficacy of biochar to remediate contaminants in soils and decrease the preparation cost, which can form economical biochars with advanced performance. Moreover, it has been found that the biochar property is mainly affected by pyrolysis conditions and raw biomass materials, so it is crucial to optimize the production systems and select appropriate raw feedstocks to generate biochar with desired properties for effective soil remediation. For example, looking for new pyrolysis methods is the recent trend, where microwaves as the heating mechanism is an emerging technique for pyrolysis and it can not only reduce the overall costs, but also improve the energy efficiency. Compared to the conventional heating method, the microwaves radiation can achieve fast pyrolysis of biomass with clean and more open porous structures, and better quality and quantity. Less formation of hazardous products and

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emission of pollutants are observed in the microwaves pyrolysis process. Apart from this, specific engineered modifications of the biochar can potentially enhance its remediation performance including biological, chemical and physical properties, and adsorption capacity. There are four common engineered methods to modify biochar, including biological modification, chemical modification, physical modification and other modification treatments (e.g., magnetic exposures). More field and full-scale applications should be conducted to evaluate the viability and feasibility of these methods in the natural environment. Furthermore, more efforts should be made to enhance the understanding of engineered biochars on their remediation potential and efficiency, and assess the impacts of their application on the soil. Although the good remediation performance of biochar for soil has been confirmed by various studies, the residence time of biochar or its aging in the soil is still a big challenge. This is because the remediation efficiency of biochar can be negatively influenced by biochar aging as well as the growth of soil microorganisms. Even though additional fresh biochar may be a possible solution, this may increase overall costs. Therefore, it is important to enhance the understanding of the aging process, which facilitates a better understanding of biochar application rate and frequency, and the long-term environmental fate of the sequestered contaminants, increasing the remediation efficiency. Simultaneously, it is needed to monitor and evaluate the long-term and short-term impacts of biochar application on soil remediation over time. Furthermore, the impacts of biochar on soil bacteria are still unclear since such impacts are associated with the biochar property and the amount of biochar added. Thus, further discussion on the intrinsic mechanism needs more effort. Most of the current studies on the impacts of biochar on the yield of plants are conducted, so how the biochar promotes the reproduction and growth of soil biomass, fungi and soil bacteria need more research in the future. More efforts should also be made to study the soil ecosystems after adding biochar since the microorganisms, plants and soil faunas work together in the soil ecosystem. Thus, the impacts of biochar application in the soil should consider there three aspects. Besides, the large-scale biochar application in the soil also needs a thorough study of the soil ecosystem.

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11 Biochar for sustainable agriculture Tao Liua, Sanjeev Kumar Awasthia, Yuwen Zhoua, Sunita Varjanib, Zengqiang Zhanga, Ashok Pandeyc,d, Huu Hao Ngoe, and Mukesh Kumar Awasthia COLLEGE OF NATURAL RESOURCES AND ENVIRONME NT , NO RT HW EST A&F UNI VE RSI TY , Y A N G L I N G , S HA A N X I PR OV I N C E , P R C H I NA b GUJARAT POL L UT I ON CONTROL BOARD, GANDHINAGAR, GUJARAT, INDIA c CENTRE FOR I NNOVAT ION AND TRANSLATIONAL RESEARCH, CSIR-INDIAN INSTI TU TE O F TO XI CO LOGY RESEARCH, LUCKNOW, INDIA d SUSTAINABILITY C LUSTER, S CHOO L O F ENGINEERING, UNIVERSITY O F PETROLEUM AND ENERGY STUDIES, DEHRADUN, INDI A f CENTRE FOR T ECHNOLOGY IN WATER AND WASTEWATER, S CHOO L O F CIVIL AND ENVIRONMENTAL E NGINEERING, UNIVERSITY O F TECHNOLOGY SYDNEY, SYDNEY, NS W, AUSTRALIA a

1. Introduction The official “sustainable agriculture” is considered as a synthesis subjects to explore the new systems of agriculture practice, which are safety and environmental protection (Semida et al., 2019). The new agriculture can be divided into two parts, including a small system and a broad process. The first method believes that the new system is enclosed, and agriculture should protect its production resources, such as maintaining soil nutrients, protecting surface water and groundwater, inventing renewable energy, and finding solutions to sustain the agricultural system of climate change. A broader goal of the second method not only divided rural and urban areas into two parts but also devoted the sustainability of the vast territory and social communities. Additionally, wastes of urban areas need agriculture to treat, such as recycling sewage sludge, developing rural economics, and providing urban residents with rural landscapes (Awasthi et al., 2018a). Before the 40 years, the beginning of the green revolution has fed a booming population, but it is generally believed it is unsustainable, would damage the environment, and cannot meet the demand for a sustainable development strategy (Barrow, 2012). According to an official investigation, agriculture is the main contributor to greenhouse gases (GHGs), which accounts for slightly less than 25% of the total global anthropogenic greenhouse gases (Smith et al., 2007). About 52.84% of global anthropogenic CH4 and NOx emissions are mainly attributable to changes in agricultural use and forestry (Smith et al., 2008). Nevertheless, the rapidly growing world population (excepted to reach 9.6 billion by 2050) will inevitably lead to an increasing in the demand for food generation, while the supply of arable land is decreasing (United Nations, 2013). Therefore, establishing effective

Current Developments in Biotechnology and Bioengineering. https://doi.org/10.1016/B978-0-323-91873-2.00009-1 Copyright © 2023 Elsevier Inc. All rights reserved.

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sustainable agriculture practices to increase crop yields while mitigate GHGs emissions have become more urgent (Qin et al., 2021). In the past few years, research on the integration of naturally derived and/or organic materials into sustainable agriculture has increased significantly (Awasthi et al., 2018b). The food generation and resistance of biotic and abiotic stresses could be utilized by some environmentally friendly biostimulants. The most potential candidate options contain humic and fulvic acids (Li et al., 2021), organomineral fertilizers (Liu et al., 2020a), and biochar (Akhtar et al., 2015; Rodriguez et al., 2021; Mona et al., 2021). Fig. 1 showed the pathways of biochar effect in soil for better crop production. These additives can effectively optimize soil fertility, accelerate plant growth, and enhance plant tolerance to adverse conditions. Moreover, the application of practical increased agricultural productivity and supported a series ecosystem services. Biochar is determined by the biomass residues of organic materials generated from the combustion process (pyrolysis) in the absence of oxygen completely or partially. Table 1 illustrated the performance of biochar addition on the sustainable agriculture. Presently, as a potentially valuable agricultural input method, biochar has received attention in improving soil fertility, helping new agricultural system production, and alleviating the adverse impacts on different biological and abiotic stresses (Xiao et al., 2017). The use of pyrolytic biomass or black carbon can be traced back at least 5000 years ago. The original inhabitants have used charcoal in agriculture. Emerging literature generally provides that the application of biochar can be a key input to maintain production while decreasing

FIG. 1 Pathways of biochar effect in soil for better crop production. indicated possible pathways.

Indicate primary pathways and

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Table 1

Feedstock

The performance of biochar addition on the sustainable agriculture. Pyrolysis temperature (°C)

Acacia green waste

550

Hardwood coniferous wood Beech wood

750

350–450

–

Utilized dosage

Performance

References

47 t ha biochar and 10 t ha 1 compost 8 t ha 1 biochar and 63 t ha 1 compost

Improve microbial abundance; change microbial structure; enhance macroscopic porosity and bioturbation effects Increase microbial abundance and activity; no effect on available copper

Abujabhah et al. (2016)

100 mg/kg biochar and 100 mg/kg compost 20 mg ha 1 biochar and 32.5 mg ha 1

Increase plant height, total organic carbon, total nitrogen content; decrease of ammonium Increase soil organic matter content, nutrient level, and water storage capacity

Schulz et al. (2013)

Increase total organic carbon and water extraction organic carbon; reduce the utilization rate of cadmium and zinc; increase the availability of Cu Improve the growth of Sesbania, Mallow, and total biomass

Liang et al. (2017b)

Improve the grain yields and N uptake

Agegnehu et al. (2016)

1

Residues of charcoal production Rice husk

500

24 g compost +16 g biochar in 400 g soil

Peanut shell

350

Acacia

350–450

0.75% biochar and 0.75% compost (wt %) 2 t ha 1 biochar, 10 t ha 1 compost and 92 kg N ha 1

Mackie et al. (2015)

Liu et al. (2012)

Luo et al. (2017)

pollution and reliance on chemical fertilizers (Wainaina et al., 2020). Recent studies have also shown that adding biochar can optimize soil fertility, soil water holding ability and nutrient absorption, which sequestering carbon and mitigation GHGs (Liu et al., 2021a). Relevant studies have shown that adding biochar to the soil can increase crop yields and relieve plant pressure related to drought. Moreover, despite the increasing benefits of biochar applications, there are also many obstacles to the utilization of biochar in sustainable agriculture (Awasthi et al., 2020a). Surely, the huge difference of biochar is our biggest concern, not only in their nutrient utilization and pH value, but also in their chemical composition and physical characteristics, depending on the nature of the raw material, pyrolysis conditions. The research progress in recent years has provided us with a scientific evaluation basis (Liu et al., 2019c). The relationship between the characteristics of biochar and its effects on soil characteristics, plant growth, yield, and resistance to biotic and abiotic stresses (Gautam et al., 2021). Among all of this book chapter is to understand a knowledgeable overview of selective references scientific publications that lecturer the different pressure on biochar in the soil, with the biochar on sustainable agriculture of reviews increasing.

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2. Evolution of biochar production technologies for sustainable agriculture purposes Generally, common wastes mainly include agricultural residues, biomass crops, manure and sludge and other wastes, which can be used as raw materials to synthesize biochar (Ravindran et al., 2021). Different raw materials have different element composition ratios, which show different properties. Therefore, biochar derived from different raw materials has different properties. Biochar and pyrogenic black carbon (Pyc) are the thermal decomposition production of biomass, which can be obtained in the presence of limited oxygen or anaerobic (Kavitha et al., 2018). Biochar provides a wide scope of agriculture advantages from enhancing the amount of organic compounds in the soil to optimize the physical, chemical and biological characteristics of the soil (Sun et al., 2015a, b). It also mitigates climate problems by storing CO2 in biochar to amend soil and reduce carbon emissions, due to its high stability under humid conditions. Biochar is hydrophobic in nature, resulting in them an attractive adsorbent for CO2 capture after combustion (Zhou et al., 2021). Due to soil fertility and abundant microbial communities under charcoal, biochar-based soil fertilizers that replicate these soil conditions are of current interest (Liu et al., 2020b). Biomass raw materials used to produce biochar, including organic materials, agriculture waste, forest residues, and manure (Kavitha et al., 2018). Generally, several thermochemical decomposition techniques can be used to produce biochar; pyrolysis, hydrothermal carbonization, gasification, and baking. Table 2 showed all kinds of feedstock and methods for biochar preparation. Pyrolysis, most commonly methods, utilized for biochar generation; containing fast pyrolysis, slow pyrolysis, flash pyrolysis, solar energy, and microwave assisted pyrolysis (Tan et al., 2017). Pyrolysis biomass included cellulose, hemicellulose and lignin which were cross-linked, depoliticized, and fragmented by unsufficient oxygen supply or a Table 2

Various of feedstock’s and methods for biochar preparation.

Feedstock

Preparation condition

Carbon content

References

Bamboo Wheat husk Straw

Pyrolysis, 500°C Pyrolysis, 500°C, 20 min Gasification 700–750°C

83.6% 50.5% 48.4%

Laminaria japonica Oil palm shell Pig manure Wood chips Peanut hull

Pyrolysis, 700°C, 1 h Pyrolysis, 700°C, 1–4 h Pyrolysis, 400°C, 1 h Gasification, 1200°C, 0.5–0.75 h Hydrothermal carbonization, 300°C, 5 h Hydrothermal carbonization, 230°C, 2 h

26.7% 82.6%–89.41% 44.13% 80.6% 56.3%

Wang et al. (2017) Kalderis et al. (2017) Hansen and Nielsen-Hauggaard (2015) Jung et al. (2016a) Hamza et al. (2016) Kołodynska et al. (2012) Wiedemeier et al. (2015) Xue et al. (2012)

71.38%

Zhou et al. (2017)

Banana peels

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completely anaerobic atmosphere at a temperature above 300°C, caused thermal decomposition to produce solids, condensable liquid and gas products. The solid powder is named biochar, at the same time the gases will be generated, containing COx and CxHx (Magdziarz et al., 2020). In traditional pyrolysis, approximately 10%–20% of biomass is transformed into biochar and 40%–75% is transformed into pyrolysis oil. The hydrothermal carbonization (HTC) method is a process similar to the natural coalification process in coal petrology. The biomass can be decomposed and carbonized in an aqueous medium at an ideal temperature and autogenous pressure (Liu and Balasubramaniam, 2014). It is different from pyrolysis that wet biomass can be directly applied to HTC, which makes it an interesting technology for thermal decomposition because biomass is propertied by its high moisture and energy cost intensive. The approaches also need relatively little energy because it has a less activation energy temperature range of 150–300°C for pyrolysis. Under the HTC, hemicellulose is first decomposed at the temperature (500°C) and anoxic conditions. The fluidity bed reactor is the most ideal reactor in practical utilization due to its high heat and mass transfer rate and uniform temperature separation. While the yield of thermal gasification biochar is lower, this technology is more stable to generate biochar than pyrolysis (Hansen and Nielsen-Hauggaard, 2015). The gasification of biomass is economically feasible, because electricity and heat generation could be used in upstream processing raw materials or downstream processing, which related the generation of biochar. Torrefaction (gentle pyrolysis) is another potential thermal decomposition project to update biomass into biochar in an inert condition at 200–300°C (Chen et al., 2017), which includes dry and wet torrefaction. Same as other thermal decomposition processes. The physical and chemical characteristics of the biochar generated rely on the torrefaction temperature and residue time. It is reported that the temperature (275 and 300°C) of 60 and 30 min will lead to the decreasing of biochar production, and the light roasting (200°C) yields about 80% of the biochar with the increasing of residue time (Chen et al., 2017). Wet approach in mainly used for the conversion of high-moisture biomass such as microalgae, which requires an energy-consuming drying process (Gan et al., 2020). As the temperature increasing, drying will cause serious degradation of cellulose, lower crystallinity, and increasing carbon residues. It is considered to be a better pretreatment method for rapid pyrolysis of biomass.

3. Modification of biochar for sustainable agriculture In order to adjust the characteristics of biochar to meet the necessary of the environment, people have to adopt a variety of methods to change biochar (Yuvaraj et al., 2021). The commonly used modification methods contain physicochemical modification. It is the

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most population utilized method that the chemical modification, which mainly contains acid, alkali, oxidant, metal salt or oxidant and carbonaceous modification. Stem and gas purification are included by physical modification. Fig. 2 illustrated the modification of biochar. The main purpose of acid modification is to eliminate impurities (i.e., metals) as acid functional groups were introduced by the biochar. Common acids contain HCL, H2SO4, HNO3, H3PO4, and C2H2O4 (Rajapaksha et al., 2016). The ash content of reed-derived biochar modified with 1 M HCL decreased from 29.5% to 11.8%, providing hydrophobic adsorption sites for pentachlorophenol (Peng et al., 2016). Modified bamboo biochar with HNO3 (2 M), and the added carboxyl, lactone, phenol, and carbonyl groups on the surface of bamboo biochar (Li et al., 2014). The surface area of biochar could be changed by the acid modification, and the type and concentration of acid also have different effects. Taking the reed biochar as an example, the surface area increased from 58.75 to 88.35 m2 g 1 after 1 M HCL treatment. After 2% H2SO4 treatment rice straw biochar, the surface area decreased from 71.35 to 56.9 m2 g 1. The chemical modification of H3PO4 method is widely utilized agent. The concentration of H2PO4 has a positive effect on the performance of biochar adsorption. Many factors affect the surface area of biochar, such as acid concentration, type, raw materials, and preparation conditions. Currently, due to the various conditions and raw materials, attaining the same results was difficult. The main purpose of basic modification is to enhance the surface area and oxygencontaining functional groups. Commonly used alkaline agents are KOH and NaOH. It is reported that NaOH reduces the surface area from 4.4 to 0.69 m2 g 1 in the wheat straw biochar. It can be considered that the impact of alkali modification on the surface area of biochar is also affected by raw materials and environmental conditions. According to the previous report, the surface area will increase by 2885 m2 g 1 used the NaOH modification significantly, which was higher than KOH modification (1940 m2 g 1) (Cazetta et al., 2011). In addition, there is no effect of the surface area by the NaOH modification of bamboo-derived biochar (Fan et al., 2010). Except for different types of alkali, the ratio

FIG. 2 Modification methods of biochar.

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of alkali also has a significant effect on the performance of biochar. When the mass ratio is 3, high-performance biochar is produced. In sum, the influence of alkali modification on the contrast surface areas depends on the raw materials and conditions. Modification with oxidants can enhance the numbers of oxygen-containing functional groups on the surface of biochar. The biochar derived of peanut shell hydrothermal carbonization was modified by H2O2 which can increase the surface of biochar the oxygencontaining functional groups, especially the carboxyl group and increase the adsorption ability of Pb. Modification with hydrogen peroxide can reduce the pH value of pine biochar from 7.16 to 5.66 and increase the oxygen-containing functional groups (Huff and Lee, 2016). While the H2O2 concentration exceeds 10%, adsorption ability of modified biochar for methylene blue is lower than that of unmodification biochar. This shows that the adsorption capacity of biochar varies with the target pollutants. Therefore, the nature of the target pollutant should be considered before using hydrogen peroxide for modification. Similar to H2O2 and KMnO4 modification can also increase oxygen-containing functional groups. Additionally, potassium permanganate modification enhances the surface area of biochar from 101 to 205 m2 g 1. The modification of biochar by KMnO4 improves the adsorption ability of pecan biochar for Pb(II), Cu(II), and Cd(II). Modification with metal salts or metal oxides can change its adsorption, catalytic, magnetic, and other properties. In order to enhance the adsorption of target pollutants, for example, the adsorption ability of biochar for anionic dyes is very low, because the surface of biochar is negatively charged. The surface characteristic of biochar can be changed by the metal modification, which improves the adsorption ability of biochar for anionic dyes. It is difficult to reused biochar from water when it is used to remove pollutants in water and wastewater, because of the small size biochar. The magnetic characteristics of biochar are increased by the iron salt or iron metal oxide modification, which is beneficial to the recycling of biochar. Improving the catalytic performance of biochar. In order to improve the catalytic performance of biochar in the per-sulfate activation system, metal salts or metal oxides can be used to compose metal-biochar compound. Iron(III) modified rice husk biochar enhances the removal rate of As(III) and As(V). Meanwhile, the modified pig manure biochar with MnO2 can improve the ability of elimination of Pb (II) and Cd (II) (Liang et al., 2017a). In order to improve the magnetic properties of large seaweed biochar, nanocomposite Fe3O4 is utilized to modify the biochar ( Jung et al., 2016b). The rice husk biochar was modified with nanoscale zero-valent iron, which will improve the catalytic performance. Liu et al. (2015) reported that iron modified the corn stalk and found that the modified biochar was stored by Fe3O4 small particles, which optimized the efficiency of phosphorous elimination in the aqueous solution.

4. Influence of biochar on soil nutrient dynamics, enzymes, and microbial community The existing literature shows that adding biochar to the soil can constitute a medium that stimulates the activity of soil microorganisms, thereby affecting the characteristics of soil

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microorganisms. The high surface area of biochar and the ability to adsorb nutrients provide a very favorable habitat for the colonization, growth and reproduction of bacteria, actinomycetes, arbuscular mycorrhizal fungi, and other microorganisms. Biochar has been found to change the composition of the microbial community, because its companion organisms increase the bulk density, pH, water, and nutrient availability in the soil matrix. As a soil amendment, biochar may have a positive effect on diazotrophs (nitrogen-fixing bacteria). Diazotrophs are free-living soil bacteria (such as Azotobacter or Azospirillum), or reciprocal bacteria associated with plants that form nitrogen-fixing nodules on roots (such as root nodules that form legumes bacteria).

4.1 Biochar and climate change It is necessary to briefly mention the function of biochar in eliminating global warming. In this process, it is worth noting that in the 2015 Pairs climate agreement, a goal was set for participating countries to hold creases in the global average temperature below 2°C higher than the pre-industrial level, and limit the temperature rise with 1.5°C of the pre-industrial level (IPCC, 2014). Although to achieve the goals of the Paris Agreement, it is necessary to adopt traditional greenhouse gas emission strategies such as reduction of fossil fuel consumption, but it is also urgent to take actions simultaneously through sustainable carbon dioxide removal (CDR) technology and planned strengthening of natural carbon sinks and reduce negative emissions (Gasser et al., 2015). In order to achieve a substantial reduction in greenhouse gas emissions, it is necessary to store carbon in terrestrial carbon sinks. Increasing the input of organic matter to the soil, or reducing the decomposition rate of soil organic matter, or the net carbon increase effect of both can increase the soil carbon storage. In this regard, the utilization of biochar in the soil has been proven to increase the net carbon of the soil, and it can also increase plant biomass production by increasing the nutrient supply to plants and improving nutrient and water use efficiency (Liu et al., 2019b). Therefore, the application of biochar in soil has been considered as a critical approach to modern global “climate-smart soil” management practices.

4.1.1 Impact on soil carbon sequestration In terms of reducing carbon dioxide emissions into the atmosphere, climate change has attracted more and more attention. Soil played a critical role in the carbon cycle in the global and as an important carbon sink to affect the climate change directly (Liu et al., 2021b). Carbon storage has been suggested as a method to reduce carbon dioxide emissions in the soil. Biochar also has higher resistance ability to biodecompose one to its highly polycondensed aromatic functional groups (Awasthi et al., 2017). Therefore, it is believed that biochar has a positive impact on soil carbon sequestration. A lot of research has been conducted on the impact of biochar on soil carbon sequestration. However, the increases and decreases of CO2 emissions have been observed, the similar results have not been obtained (Sheng and Zhu, 2018; Yousaf et al., 2017). For example, fire carbon amendment soil could promote the transformation of organic

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carbon. Yousaf et al. (2017) reported that adding wood chips-derived biochar to the soil will reduce carbon mineralization, leading to greater carbon sequestration. Carbon content of biochar prepared from crop residues and wood materials is higher than that of biochar prepared from manure sources. Carbon content of waste biochar has a wide range (19.2%–84.0%), indicating that there are differences in its initial composition. The carbon fixation effect of biochar comes from the carbon of biochar, not the carbon from soil organic matter. Some researches published some reviews article which studies and made meta-analysis of degradation and main impacts. They observed that the degree of mineralization of organic matter in the fertility soil is higher than that of high-fertility soil before adding biochar. Meanwhile, soils with low organic content are more carbon mineralized than soil with high organic matter concentration. Additionally, the cultivation time has a significant impact on the start-up effect of biochar. The carbon of biochar can be divided into two parts, included soluble carbon and insoluble carbon. After adding biochar to the soil, the soluble carbon is easily used by soil microorganisms, resulting in increased initial carbon mineralization. This explains the phenomenon of adding biochar to promote carbon mineralization. Factually, the content of insoluble carbon in biochar is much higher than that of easily soluble carbon (Crombie et al., 2015). These stubborn carbon can last a long time in the soil. It can be seen that the carbon input gave rise to by the amendment of biochar is higher than the carbon output caused by the mineralization of volatile carbon. Generally, the impact of adding biochar on carbon sequestration is still unclear. The starting effect of biochar under different raw material types and pyrolysis conditions is different. It is very necessary to study the correlation between the starting effect and the type of raw material. It is considered that pyrolysis conditions have a great potential effect on the physical and chemical characteristics of biochar, and also critical to understand the correlation between pyrolysis conditions and the carbon fixation effect of biochar. In addition, when studying carbon sequestration induced by biochar, soil composition should also be considered.

4.1.2 Impact on soil N2O and CH4 emission With atmosphere, N2O is the most important ozone-depleting compound. Before industrialization, the N2O concentration in the atmosphere was 270 ppb, while the current atmospheric abundance is about 324 ppb (Ussiri and Lal, 2013). Widespread use of nitrogen fertilizer is the main source of global man-made N2O emissions. Emission of N2O is the primary greenhouse effect potential, which generated via conversion of nitrogen in the soil, while N2O is generated through non-biological redox reactions. Rondon first reported the reduction of N2O emissions after biochar modification. Soybeans grown in the oxidized barren soil of the Colombian savanna have reduced N2O emissions by 50%, and grass N2O emissions have been reduced by 80%. This hypothesis indicates that the impact of biochar on N2O emissions contains abiotic and biological mechanisms, such as biochar liming impact, interaction with nitrogen, interaction with dissolved organic carbon, effect on soil aeration and volatilization of toxic compounds. Cayuela et al. (2013a) determined a meta-analysis of 261 experimental treatments between 2007 and 2013. They observed that

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laboratory and field studies N2O emission would reduce 54%. The datum comes from 88 publications from 608 observations as of May 2016. The overall reduction in N2O is about 38%, but the reduction in N2O emissions after 1 year is almost negligible. In addition, biochar has the best N2O reduction effect on sandy soil and paddy soil (Borchard et al., 2018). Cayuela et al. (2013b) conducted in-depth research on N2O emission reduction and emission reduction mechanism. They observed that biochar greatly influence nitrogen recycling of denitrification, reducing N2O emissions by 10%–90% in 14 different soil. Changes of soil pH value caused by the buffering capacity of biochar are an important aspect of reducing N2O emissions. In addition, biochar promotes the electronic transfer of denitrifying microorganisms, promotes the reduction of N2O, and exerts a lime effect. They also found that adding of biochar could reduce the emissions of N2O and N2 in all cases, but did not reduce the ratio of N2O/(N2 + N2O). Biochar as a soil amendment can affect the nitrogen cycle. The abundance of microbial nitrogen cycle genes will be increased by the addition of activated switch grass biochar in the subsoil of arid areas (Hale et al., 2012). The addition of biochar promotes nitrification and denitrification. Through redundancy analysis, the chemical properties of the soil have changed, leading to changes in the soil microbial community, thereby regulating soil N2O emissions and nitrogen cycle. In addition to reducing N2O emissions from improving soil, biochar can also reduce soil N2O emissions. Adding a small amount of biochar in the compost with soil incorporated can promote the nitrogen cycle by increasing the content of NO3 -N, indicating higher nitrification activity. With the large influx of nitrogen in urine, the ratio of denitrifying agent to nitrify agent increases. This is correlated to the control of different N2O production pathways in the soil, and biochar has diverse effects on these processes. The causes of these phenomena and the actual duration of mitigation effects should be further studied. The concentration of CH4 in the Earth’s atmosphere has increased. Since 1750, the contribution rate of global warming has been 150%, accounting for 20% of the global warming effect (Korzh et al., 2011). The emission of CH4 is emitted through natural resources, including wetlands and human activities. In order to assess the real benefits of biochar in reducing greenhouse gas emissions, it is necessary to quantify the impact of biochar on improving soil CH4, particularly in wetlands, where the soil is often drained and flooded, thereby facilitation CH4 and N2O emissions. Dong et al. (2013) compared the response of biochar and straw utilization to CH4 emissions in a paddy field experiment. In a two-year revision, the results showed that rice straw-derived biochar is more effective than bamboo-derived biochar in reducing CH4 emissions from rice fields. Compared with the direct return of straw to the field, straw biochar reduced CH4 emissions by 47.30% to 86.43% during the rice growth cycle. This is because the activity of methanogens decreases with the increasing of pmoA gene abundance and CH4 oxidation activity. Through 4 years of field trials, the impact of biochar improvement on greenhouse gas emissions was studied (Wang et al., 2018). Straw-derived biochar has reduced the total annual CH4 emissions by 20%–51% in 4 years. The results show that the improvement of 24 t ha 1 biochar is a

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sustainable, effective, and economical GHGs emission reduction measure. Mohammadi et al. (2015) has calculated the impact assessment of the open burning of residues and the use of life cycle conversion into biochar on climate change. In two systems, the largest contributor to the carbon footprint of rice is CH4 emissions. Compared with open burning of straw, biochar treatment reduced the carbon footprint of summer rice and spring rice by 14% and 26%, respectively. After 8 years of biochar improvement, the value has increased to 38% and 49%, respectively. These evidences suggest that the increase in carbon dioxide emissions is solely due to the release of nonbiological CO2 from biochar. The average inhibition rate of biochar to improve soil N2O is 63%, which is also related to the initial N2O production. This kind of biochar has a predictable impact, although the type of soil is different, the greenhouse gas emissions are also different.

4.2 Role of biochar in soil bioremediation 4.2.1 Remediation of heavy metals Soil heavy metal pollution refers to human activities lead to excessive deposition of trace elements in the soil, causing the content to exceed the background value. Since heavy metals are not easily decomposed by soil microorganisms, they tend to accumulate in the soil, and even transform into more toxic methyl compounds, which may accumulate in human body through the food chain due to the crops planted in soil, thus posing a serious threat to human health. Biochar, as an environmentally friendly and efficient adsorption material, is a new type of material for sustainable development. It has a large specific surface area, higher pH and the property of adsorption. The higher pH can reduce the mobility of heavy metals, which can increase the sequestration of heavy metals in the soil. And its surface has oxygen-containing functional groups such as carboxyl groups and hydroxyl groups, which can increase the cation exchange in the soil and improve the stability of heavy metal in the soil. Table 3 lists the biochar as well as biochar composites used in relevant studies and the remediation effects of these materials on specific pollutants, from which it can be seen that biochar and its composites can play a role in the treatment of heavy metal pollution. Biochar can directly adsorb heavy metal ions through physical adsorption, electrostatic adsorption, ion exchange, complexation, precipitation, and redox by utilizing its properties of high specific surface area, CEC, organic carbon content, and active functional groups. (1) Physical adsorption. Biochar can adsorb heavy metal on its surface or diffuse them into pore channels, and immobilize them through van der Waals forces and other effects. Usually, the large specific surface area and porosity ease the physical adsorption of heavy metals by biochar. Deng et al. (2019) studied the adsorption capacity of biochar made from rice straw at 400 and 700°C for Cd and Ni, the results showed that the biochar made at 700°C had a larger specific surface area and higher adsorption capacity for heavy metals. (2) Electrostatic adsorption. Biochar can immobilize heavy metals through using its surface charge to adsorb heavy metals through electrostatic interaction. The electrostatic adsorption of heavy metals by biochar depends on the environmental pH and

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Table 3 Remediation of heavy metal pollutants in soil by biochar and its composite materials. Biomass and its composite material

Pyrolysis temperature (°C)

Wheat stalk

550

Bagasse uptake nanohydroxyapatite Wheat straw

600

Poultry litter uptake Fe3O4 Peanut stalk

500 500

Rice husk uptake sulfur

550

Rice straw uptake iron

300

Biochar-manganese oxides composite Cotton peanut stalk

600 500

Cow dung rice straw

350

450

Field Paddy soil

Paddy soil Paddy soil Paddy soil

Paddy soil Red soil Brown soil Farmland

Contaminants

Adsorption effect

References

1

Cd

1.04 mg kg

Pb

460.52 mg kg

Cd

21.84 mg g

Pb

689.64 mg kg

1

Pb

11,398 mg kg

1

Hg

67.11 mg g

As

22.03–43.06 mg kg

Cu

1.91 mg g

Cu

13.56 mg g

1

Cd

10.8 mg kg

1

1

1

1

1

1

Bian et al. (2014) Yang et al. (2016) Cui et al. (2016) Lu et al. (2017) Xu et al. (2018) Wu et al. (2018) Wu et al. (2018) Yu et al. (2014) Jin et al. (2016) Cheng et al. (2017)

the zero charge point (pHPZC) of biochar. When the pH of the medium is >pHPZC, the surface of biochar is negatively charged and can electrostatically adsorb with cationic heavy metals; when the pH of the medium is 93%); in agricultural soils, the biochar produced at 700°C showed better adsorption effects on Pb and Zn. The ambient temperature at the time of biochar application can also affect its adsorption capacity. Arabyar et al. (2017) prepared a biochar-clay-chitosan composite, which had a maximum adsorption capacity at 25°C, which is much higher than that of the original biochar. Meanwhile, type, concentration, and morphology of heavy metal can lead to different results of biochar for soil remediation. The study by Khan et al. (2014) showed that biochar significantly reduced the disaffected state concentrations of Pb and Cd in the field, but had no effect on Zn, which may be due to the different activities of different heavy metals. In complex heavy metal contaminated soil, the competing adsorption of multiple heavy metals may make one heavy metal more active, e.g., Pb forms complexes more easily on the surface of biochar than Cd, while manganese oxides sorbed Pb significantly better than Cd. Zheng et al. (2015) showed that the application of biochar resulted in a significant reduction of Cd, Zn and Pb in the effective state, while As increased.

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Biochar promoted the reduction of As(V) to As(III), resulting in increased as toxicity and mobility in contaminated fields. Therefore, different geochemical behaviors of cationic heavy metals such as Cd and As make it difficult for them to be immobilized. Biochar can also reduce the accumulation of heavy metals in plants by reducing their mobility and bioeffectiveness in the soil. In many field trials, biochar was effective in reducing the accumulation of heavy metals in crops, but there were exceptions. A study by Zheng et al. (2015) showed that the application of beanstalk and rice straw biochar resulted in a significant reduction in Cd content in rice, but the difference in Zn contented was not significant. This may be due to the fact that rice plants have a large number of Zn transporters dedicated to Zn uptake and transport, while biochar addition did not significantly limit the elemental transport. The accumulation of heavy metals in plants was also related to the pollution status. Zhang et al. (2016) found that the concentration of lettuce tissues did not decrease after biochar treatment of heavily contaminated soil with Cd, probably because the soil organic matter concentration was too high and the amount of biochar applied was not sufficient to fix the organic matter in the soil. Although biochar has good physicochemical properties and has achieved some success in research on environmental remediation, its specific surface area, type and number of surface functional groups are still deficient and its adsorption capacity is small in some extent. At present, chemical, physical, and biological methods are often used to improve the physicochemical properties of biochar, and the adsorption capacity of modified biochar has been significantly improved, and modified biochar has become a hot spot for research in soil heavy metal remediation. Different modified biochars are used in the remediation of heavy metal contamination in the land with different characteristics, and the physicochemical properties of biochar are adjusted to modification to achieve the best remediation effect. For example, chitosan modified biochar has stronger adsorption properties for lead, copper, and cadmium in soil than unmodified type. Zhang et al. (2016) made phosphorus modified biochar by adding potassium phosphate to cow dung biochar and applied it into the soil contaminated with lead and chromium complex for the test, and found that the modified biochar promoted the conversion to weak acid extracted state of lead and chromium to residue state and improve the quality of soil. Loading nano zero-valent ironed on the surface of biochar promoted the reduction of hexavalent chromium and reduced the biomoldability of chromium in the soil. It can be seen that the remediation mechanism of modified carbon and biochar for heavy metal contaminated soil is monoclinic, but modified carbon has better remediation effect and characteristics, which is more suitable for the study of soil heavy metal contamination remediation.

4.2.2 Remediation of organic pollutants At present, there are relatively more studies on the application of biochar materials in the remediation of heavy metal contaminated soils, while there are relatively fewer studies on their application for the remediation of contaminated soils, but the trend is increasing year by year, mainly focusing on pesticides, aromatic compounds, antibiotics and other organic types of pollution. The practical application of organic pollutants is

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Table 4 Remediation of organic carbon contaminated soil by biochar and its composite materials. Biomass and its composite material

Pyrolysis temperature (°C)

Field

Contaminants

Adsorption effect

References

1

Cotton wood

600

Methylene blue

174 mg g

Corn stalks uptake Fe3O4

400

Crystal violet

349.40 mg g

Wheat straw

350–650

PAHs

0.29 μg g

Olive Clay Pine, spruce

550 700

0.59 mg g 9 mg g 1 100%

Corn stalks

700

Metalaxyl tebuconazole Octadecae octadecanoic acid Atrazine

139.33 mg g

1

Bagasse

550

Latosols

Norfloxacin

620.8 mg kg

1

Cotton peanut stalk

300

Cu

13.56 mg g

1

Rice straw

500

Brown soil Paddy field

Phenanthrene

0.160 mg g

1

Loamy sand Sandy loam soil Sandy soil

1

1

1

Zhang et al. (2012) Sun et al. (2015a, b) Kusmierz et al. (2016) Gamiz et al. (2016) Hallin et al. (2017) Yang et al. (2018) Arabyar et al. (2017) Jin et al. (2016) Ding et al. (2021)

shown in Table 4. Biochar remediation of organic contaminated soil is mainly realized by adsorption, and the adsorption mechanism mainly includes three ways: distribution, surface adsorption and pore retention (Liu et al., 2020c). The increase in organic pollutants in soil in recent years has made soil remediation complicated and the difficulty of adsorption has increased, and several adsorption methods are generally required to work together. (1) Distribution effect Distribution effect is the distribution of organic pollutants between hydrophilic and hydrophobic phases according to similarity-intermiscibility theory, and the process is influenced by the organic matter content of the soil. Distribution effected to occurs mainly to the noncarbonized fraction of biochar, which is a monolinear sorption process of nonionic organic compounds in the soil (Arabyar et al., 2017). Distribution effect is determined by the matching and effectiveness of biochar with organic pollutants, which is high if the polarity of biochar matches that of organic pollutants; effectiveness refers to the effective solubilization of organic pollutants by amorphous organic carbon on the one hand, and the noncompetitive partitioning of high concentrations of organic pollutants on the surface of biochar on the other. Generally, biochar prepared by hightemperature cracking can adsorb neutral organic matter, while biochar prepared by low-temperature cracking has a part of the organic phase through partitioning in addition

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to surface adsorption. The smaller the H/C ratio, the more aromatic the biochar is; while the larger the (N + O)/C ratio. (2) Surface adsorption The surface of biochar has a rich pore structure, which is stable and has a large specific surface area. This enables the surface of biochar to adsorb organic pollutants and make the organic pollutants interact with the molecules on the surface of biochar, resulting in the enrichment of adsorbent on the surface adsorption sites, so biochar has a superb adsorption capacity. The surface adsorption process is influenced by the electronegativity, acidity, alkalinity, and aromaticity of the biochar surface as well as the nature of the pollutants. Biochar prepared by high temperature cracking (above 500°C) is characterized by high specific surface area, low polarity, and rich aromatic structure, which can interact with organic pollutants and increase the nonlinearity of adsorption isotherm. The surface adsorption of biochar can be divided into physical adsorption and chemisorption according to the difference of interaction forces between the adsorbent and the adsorbate, and when the change of thermodynamic surface free energy of adsorption is less than 40 kJ/mols, physical adsorption is dominant, and vice versa (Wu et al., 2018). The most common physical adsorption is electrostatic adsorption, which is based on the principle of weak interaction between oxygen-containing functional groups of the surface of biochar and organic pollutants; while chemisorption will have the formation of chemical bonds (such as hydrogen bonds, ionic dipole bonds, coordination bonds, or π-π bonds, etc.) or strong intermolecular interactions. (3) Pore retention There are many types of microporous structures of the surface of biochar, and few larger pore levels, so biochar is a nonhomogeneous porous solid material (Arabyar et al., 2017). Too large or too small biochar pores will all limit the retention of organic pollutants; too large pores will have an impact on the retention of small molecule organic pollutants; too small pores will make it difficult for large molecule organic pollutants to enter the interior of biochar. The study on the adsorption of naphthalene by corn straw biochar showed that all obtained by characterizing as the intraparticle diffusion model was multicollinear, and the adsorption of naphthalene by pore filling was very effective. Pore retention is not a simple physical capture, but hydrophobicity within the micropores also plays an important role in pore retention (Xu et al., 2018). It was found that the adsorption of estradiol by biochar was influenced by the pore structure, and it was hypothesized that estradiol mainly forms hydrogen bonds and π-π bonds with intrapore groups and interacts with EDA, etc. (Zhang et al., 2017). Therefore, the pore retention mechanism is very important to the adsorption of organic pollutants by biochar, and the process is related to the pore size, the composition of functional groups within the pores and the nature of organic pollutants. Because biochar has more microporous structures, it will slowly adsorb organic pollutants. Organic pollutants can only enter the interior of the rigid structure of biochar by slow diffusion or distribution and be adsorbed

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and fixed, so that the organic pollutants retained by the pores are called residual state organic pollutants (Zhang et al., 2021). (4) Combined action The mechanism of organic pollutant adsorption by biochar can hardly be explained by single distribution effect, surface adsorption, or pore retention. Different biochar properties, organic pollutant properties and adsorption environments can cause differences in the process of organic pollutant adsorption by biochar, which is generally explained by one mechanism as the main mechanism and several other mechanisms as supplementary mechanisms (Zhang et al., 2017). The analysis of p-nitrophenol (PNP) adsorption on organic bentonite revealed that a combination of partitioning and surface adsorption is required to explain the adsorption process. A study added PNP to soil and compared the changes in organic pollutant adsorption isotherms before and after the addition, and concluded that the adsorption process of organic pollutants is a combination of distribution effect and surface adsorption (Kusmierz et al., 2016). The factors affecting the adsorption capacity of biochar for organic pollutants mainly include the physical and chemical properties of biochar itself and the nature of organic pollutants. Different biochars and their composites have different selectivity and adsorption effects on organic pollutants. External factors (temperature, type of composite material, etc.) also have an important influence on the actual effect of biochar. In practical application, suitable materials and preparation methods need to be selected according to the conditions of use. Table 4 lists some of the major pesticide components and some common organic pollutants, indicating that biochar and its composites have positive effects on remediation of organic pollutants. The important physicochemical properties that affect the adsorption capacity of biochar include pore structure, elemental composition, specific surface area, aromaticity, pH, and C/N. Different raw material types, preparation conditions, and modification conditions can produce biochar with different physicochemical properties. The physicochemical properties of biochar are influenced by the lignin content of the raw materials (Lei et al., 2020). The aromaticity, stability, and carbon content of corn straw biochar are higher than those of wheat straw biochar, while its polarity and ash composition are lower than those of wheat straw biochar. The results showed that the specific surface area, aromaticity, pH, and C/N of the wood chip biochar were higher than those of the dry cow dung biochar, and the ash composition was less than that of the dry cow dung biochar, which was beneficial for improving the water holding capacity of the soil and increasing the surface adsorption of the biochar. Comparing the carbon composition of different biochars, the results showed that the stability of biochar had a positive correlation between O/C, which was mainly related to the efficiency of carbon conversion during preparation (Hallin et al., 2017). The physicochemical properties of biochar during aging also change from different pyrolysis temperatures and raw materials. Aged low-temperature pyrolysis biochar is characterized by small aromatic carbon content, large alkyl carbon content and strong polarity, and with the dissociation of functional groups of the surface of biochar, its

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alkalinity increases and CEC increases, which has an impact on the adsorption process of biochar. The specific surface area of biochar can be increased by physical, chemical and biological modification methods to achieve better adsorption. Biochar adsorption is also influenced by the polarity, hydrophobicity, aromaticity, molecular size, and other factors of organic pollutants. Biochar has aromatic structure, which can interact with organic substances with benzene ring structure and generate π-π force, so that organic substances are strongly adsorbed on the surface of biochar (Lu et al., 2017). Pollutants with different molecular sizes vary from their adsorption mechanisms and adsorption effects due to the differences in effective contact and retention effectiveness on biochar under the action of spatial site resistance. Therefore, for organic pollutants with different molecular volumes and polarities, the adsorption mechanism of biochar will change. For hydrophobic organic pollutants, hydrophobic partitioning is the main adsorption mechanism, while for strongly polar small-molecule organic compounds, adsorption of polar functional groups on the surface of biochar is the main adsorption mechanism, and the different molecular sizes of organic pollutants and the different filling effects of their micropores will also cause the adsorption amount to change. The adsorption environment also affects the adsorption of organic pollutants by biochar, which is mainly the ambient pH, environmental media and coexisting ions, etc. And most of the existing soil contamination is compound (Manna and Singh, 2019). It has been shown that in composite pollution, the adsorption strength of biochar decreases from each organic species in the case of coadsorption of several organic species, indicating that there is competition between organic species in biochar adsorption and the coexistence of organic species affects the process of organic pollutant adsorption by biochar. Biochar remediation also affects other environmental behaviors of organic pollutants in the soil. (1) Effects on volatile organic compounds (VOCs). Volatile organic matter refers to organic matter with saturated vapor pressure larger than 70. 91 Pa or boiling point less than 260°C at room temperature. VOCs usually exist on soil in the form of gas, liquid and solid, and are cumulative, hidden, volatile and toxic, posing a serious threat to the health of the public and the reuse of land. When biochar remediates soil, it can sequester organic pollutants with volatility or semivolatility in the soil and reduce their volatilization. Related studies have shown that the addition of biochar to soil significantly reduces the volatilization of hexachlorobenzene (HCB), and the volatilization of HCB decreases from the increase in biochar content (Yu et al., 2009). The results of the study on the effect of biochar on the volatilization of low-chlorobenzene in soil showed that the addition of 0.1% biochar to the soil reduced the volatilization of 1,2,4-TCB by 90%; increasing the chlorine content decreased both the volatilization reduction and the volatilization rate. The small volume and low chlorine chlorobenzene molecules are volatile and can easily enter the micropores of biochar, makes biochar more capable of adsorption of low chlorine chlorobenzene (Manna and Singh, 2019). The addition of 1% biochar to soil treated with low concentrations of 1,3-dichloropropene (DCP) resulted in a reduction in its entry into the gas phase, and biochar reduced its entry into the gas phase by nearly 50% when treating high concentrations of DCP, leaving the DCP in the soil-water system. However, the

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reduced DCP in the gas phase was not transferred to the aqueous phase, and the DCP concentration in the soil solution would become lower, indicating that the adsorption of biochar caused more DCP to remain on the soil. In conclusion, biochar can reduce the volatile effect of volatile organic pollutants and reduce their pollution to the environment. (2) Effect on the biological effectiveness of organic pollutants in soil. Bioeffectiveness is the accessibility of chemical substances to living organisms and their potential toxicity to living organisms. The ability of chemical substances to enter living organisms is related to the type of organisms, the nature of media ( Jones et al., 2011). The addition of biochar reduces the biological effectiveness of contaminants in the soil. Biochar reduces the concentration of contaminants in soil solution and their biological effectiveness by shifting organic contaminants in the soil from easy-to-desorb sites to difficult-to-desorb sites. Microorganisms can degrade organic pollutants in soil particles and soil solution and play an important role in the biological effectiveness of pollutants, while adding biochar will reduce the availability of microorganisms to certain organic pollutants and prolong their retention time of the soil (Kamau et al., 2019). Effect on catalytic degradation of organic pollutants in soil. Besides acting as an adsorbent, biochar can also act as a catalyst and catalyst carrier to catalyze the hydrolysis of some pollutants in soil. Cooxidation of biochar and hydrogen peroxide can remove organic pollutants, even some substances that are not biodegradable (Liu et al., 2019a). The biochar-soil system is complex, which will change the soil pH and catalyze or inhibit the hydrolysis of certain organic pollutants in the soil. Biochar in soil is more likely to adsorb organic pollutants, and the hydrolysis of organic pollutants in soil is affected by soil pH and other substances or components in soil, such as dissolved organic matter and hydrolytic enzymes. Biochar will accelerate the oxidativereductive degradation process of organic pollutants, which has catalytic properties similar to other black carbon materials (e.g., graphite, activated carbon) for oxidative-reductive degradation of herbicides (dimethoate and flutriafol) by dithiothreitol. The catalytic effect of biochar is mainly due to the graphitic structure of biochar which can provide sites for adsorption and oxidation-reduction, and its surface functional groups can perform catalytic, oxidation-reduction reactions to activate metals, which have catalytic degradation effects on nitro-containing herbicides.

4.2.3 Remediation of pesticide residue With the continuous increase in the world’s population, human demanded for food has also greatly increased. However, agriculture is always adversely affected by pests, resulting in from to 45% of crops being damaged every year. Therefore, the pesticides are commonly used to improve the quality of crops in agriculture. In order to increase the yield of crops, people use a lot of chemical fertilizers. Due to the low awareness of environmental protection and the lack of proper management, overuse of pesticides and fertilizers causes more than 99% of the residues to remain in soil, resulting in significant acidification of the soil, reduction of soil nutrients (nitrogen, phosphorus, potassium, etc.) with the increase of pollution, and reduction of soil void space, resulting in soil structural consolidation. Overuse of pesticides also causes harm to soil organisms.

Chapter 11 • Biochar for sustainable agriculture 319

The ability of biochar to adsorb pesticides from soil depends on its unique physicochemical properties. Biochar has more surface functional groups and large specific surface area, contains oxygen-containing functional groups such as phenolic hydroxyl and carboxyl groups, and has a highly aromatized structure, which make it possess good adsorption properties. The overall removal mechanisms of biochar for organic pesticide include: (i) direct adsorption; (ii) activated biometabolism; and (iii) catalytic degradation. In fact, the removal of organic pollutants by biochar is usually the result of the combined effect of several mechanisms, and its control of environmental pollution is mostly attributed to its modification of the migration and transformation pathways and degradation patterns of organic pollutants in soil through the above-mentioned mechanisms. 1) Direct adsorption. Related studies have shown that when biochar is added to pesticide contaminated soils, organic pesticides can reduce their risk of leaching by liquid water dissolution and runoff diffusion through massive adsorption by biochar, such as simazine, imazamox, 3,5,6-trichloro-2-hydroxypyridine (chlorpyrifos metabolite), glyphosate, etc. (Lei et al., 2020; Manna and Singh, 2019; Jones et al., 2011; Hagner et al., 2015). In the study of chlorpyrifos- and furadan-containing soils, Yu et al. (2009) found that the levels of both insecticides in shallots were reduced in soils with 1% biochar compared with no biochar treatment, by 10% for chlorpyrifos and 25% for furan, respectively. While in a study of diquat and isoxaflutole, the efficacy of both herbicides was reduced after the introduction of biochar into the soil (Xu et al., 2008). These studies all suggest that biochar can relatively weaken the physiological activity of organic pollutants by reducing their content of target organisms, which is one explanation for the reduced biological effectiveness, but there are no strong studies to confirm whether biochar reduces plant uptake by reducing the dissolved state concentration of herbicides in free soil water. 2) Activation of biometabolism. One explanation for the activation of biological metabolism by biochar is that the soluble nutrients present in biochar stimulated the growth and proliferation of microorganisms in a certain area. Another explanation is the adsorption of organic pollutants by biochar enriches a single species in a certain area, which also has a selective effect on the growth and proliferation of microorganisms, retaining the microorganisms that can grow normally in the presence of such pollutants and even use them as a nutrient source. This has a selective effect on the growth and proliferation of microorganisms, retaining populations that can grow normally in the presence of such pollutants and even use them as nutrient sources. The results showed that the addition of biochar reduced the photolysis and hydrolysis rates of the two organic pollutants to some extent (probably related to adsorption), but the microbial degradation was significantly enhanced over a longer time scale, resulted in a decrease in the half-life of methomyl. However, microbial degradation did increase significantly from longer time scales, reducing the half-life of methomyl from 40.7 to 25.6 days and that of imidazole ethacrynic acid from 43.3 to 26.5 days (Zhang et al., 2021). In addition, it should be noted that different the direct adsorption, biochar

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should be applied in appropriate amounts to activate the biometabolic activity of biochar, and excessive application may cause chemical stress to microorganisms. 3) Catalytic degradation. The catalytic degradation of organic pollutants mediated by biochar as a catalyst is the latest and hottest research point of biochar remediation mechanism in recent years, and the main research ideas are broadly divided into two aspects: photocatalytic degradation and advanced oxidation. Photocatalytic degradation refers to the use of light radiation energy to provide energy to the catalyst (biochar), which causes the electrons of the catalyst surface groups to leap to the excited state, and then the high-energy particles directly reduce the organic compounds adsorbed on the catalyst surface, or transfer energy to the oxygencontaining group on the catalyst surface to generate oxygen-containing radicals, triggering the free radical reaction to organic compounds, i.e., biochar acts as a photocatalysis in the photocatalytic degradation reaction. The biochar acts as a photosensitizer in the photocatalytic degradation reaction (Liu et al., 2019a). The advanced oxidation method can directly or indirectly convert pollutants into H2O and CO2, which has unique advantages in treating organic pollutants that are difficult to be oxidized by general oxidants and difficult to be degraded by organisms (Kumar et al., 2017). The adsorption of pesticides by biochar is influenced by many aspects, such as the type of biochar, the type of pesticide, the pyrolysis temperature, and the amount of biochar applied. Residual pesticides not only have an inhibitory effect on the growth of ground crops, but also affect the community diversity and population distribution of soil fauna. The number of soil animal taxa, the number of individuals and the community diversity decreased from the increase in pesticide concentration. As an ideal indicator organism for agrochemical contamination in soil ecosystems, earthworms are more sensitive to certain contaminants than many other soil animals and are widely used to determine the biological effectiveness of pesticide residues. One study found that the bioenrichment factor of earthworms for chlorobenzenes was significantly reduced by adding biochar to the soil, and the degradation effect of biochar was more pronounced the longer the aging time. This experiment demonstrated that biochar can reduce the biological effectiveness of residual pesticides in soil, and thus reduce the toxic effects of residual pesticides on soil animals. Besides, biochar has a large amount of oxygen-containing functions such as – COOH and –OH on its surface, which presents hydrophilic and hydrophobic properties and can maintain soil moisture to a certain extent, improve soil physicochemical properties, and provide a better living environment for soil organisms. Therefore, the application of biochar in soil can reduce the toxicity of pesticides in the short term and increase the species and number of soil organisms, but it only plays a role of adsorption, reducing its short-term risk, the potential environmental risk of residual pesticides still exists, and high concentrations of residual pesticides still need microbial degradation in the end. Soil microorganisms are the driving force of material transformation in soil, which plays an important role in the biodegradation of residual pesticides in soil. Biochar

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changes the physicochemical properties of soil by altering soil fertility, moisture, pH, CEC, EC, etc. (Duan et al., 2022). Biochar can improve nitrification and inhibit denitrification, promote the assimilation and fixation of nitrogen by microorganisms, and facilitate the accumulation of nitrate nitrogen in soil. In addition, biochar provides a good shelter for microorganisms from predation or loss due to changes in hydraulic conditions because of its large surface area and porosity, thus effectively increased the number of microorganisms in the soil (Duan et al., 2021a). Biochar provides a carbon source and various nutrients for the growth of microorganisms to ensure their normal life activities, which in turn promote the degradation of pesticide residues in soil by microorganisms (Awasthi et al., 2020c). However, the adsorption effect of biochar not only acts on microorganisms, but also strongly adsorbs organic pollutants in the soil. The desorption efficiency of diquat in biochar-applied soil decreases in increasing adsorption time. By adsorption of organic pollutants, biochar “isolates” pesticide residues from soil microorganisms and enriches them in the rich pore structure of the biochar. However, biochar reduces the availability of pesticides to microorganisms of a certain extent and enhances the desorption hysteresis, which prolongs the half-life of pesticide residues.

5. Impact of biochar on crop growth and yield Biochar is considered to be a good soil amendment. As the amount of biochar added increases, the soil capacity decreases, the density decreases, and the total porosity of the soil increases. The greater the amount of biochar added, the smaller the saturated hydraulic conductivity of the soil, which has a positive effect on enhancing water retention performance of sandy loam soil (Duan et al., 2021b). The addition of biochar not only has an impact on soil nutrients, but also stabilizes carbon sequestered in the soil carbon pool, thus playing a role in carbon sequestration and emission reduction. Biochar can also significantly increase the organic carbon content of topsoil. In addition to improving soil carbon storage, biochar application for agricultural fields can also increase phosphorussolubilizing bacterial abundance and influence phosphatase production, thus increasing the content of available phosphorus, and ultimately leading to a significant increase in soil fertility and crop productivity, and the total soil carbon content, microbial mass carbon and bioavailable phosphorus are most affected by biochar application. Meanwhile, biochar can reduce the heavy metal concentration on plants by immobilizing heavy metals in contaminated soil and reduce toxicity to improve the productivity of rice fields (Duan et al., 2020). The addition of 1.5 and 3.0 t/hm2 of sludge biochar increased grain yield by 1.1 and 1.8 times, respectively. Meanwhile, biochar can improve soil water holding capacity and increase the content of nutrients and micronutrients. Biochar could also significantly enhance soil pH and organic matter content, which increased maize yielded by 18.92%–27.67%. Biochar could also change soil microbial and soil enzyme activities and improve the root zone environment for crop growth, thus promoting crop yield increase (Duan et al., 2019). Meanwhile, biochar could inhibit soil fungal growth,

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maintain soil bacterial community structure, and increase the ratio of bacterial to fungal content, resulting in 11.4%–26.8% yield increase in cucumber; biochar could improve soil enzyme structure and reduce soil heavy metal activity, thus increasing tomato yield. Biochar can also be applied to the soil as an organic fertilizer carrier, thus promoting crop growth. The study by Ibrahim et al. (2020) showed that compared with conventional fertilizer treatments, biochar replacing 10% of chemical nitrogen fertilizer could increase the yield of pod pepper by 7.3%; the nitrogen fertilizer utilization rate increased by 4.45 percentage points and significantly improved the quality of pod peppers. However, there is a risk of yield reduction if the proportion of biochar replacing chemical N fertilizer is too high. In a long-term soybean trial conducted by Zhu et al. (2019) showed that biochar application increased yield by 17.9%–33.3% compared with the control; under the winter wheat-summer corn crop rotation system, biochar dosage had a significant effect on yield in both seasons, with yield increases of 11.3%–58.3% (winter wheat season) and 6.9%– 12.6% (summer corn season), respectively; compared with the nitrogen fertilizer-only treatment, biochar with nitrogen fertilizer increased yield. Biochar with N fertilizer increased vegetable yielded and improved N fertilizer utilization compared with N fertilizer only treatment. The addition of biochar significantly increased the yield of celery by 30.3%–31.6% compared with no biochar addition, but the effect on eggplant was not significant. Yuan et al. (2021) found that biochar prepared for different raw materials also showed significant differences in crop yield enhancement, with the average yield increase of 16.5% for livestock manure-based waste, which was higher than that of biochar prepared for other types of raw materials. The effect of biochar application for crop yield in a single season or in the short term was not significant, but in the long term, biochar could increase crop yield by 6.5%; making biochar into biochar-based fertilizer significantly increased the yield of corn and rice compared with traditional compound fertilizer, and the right amount of biochar could stabilize or improve crop yielded. Wang et al. (2020) found that biochar had the effect of promoting spinach yield increase, and the spinach yield increase was 68.74%–214.94%, and the crop yield were positively correlated with the amount of biochar added. Kamau et al. (2019) studied that corn seed yielded increased from 0.9 to 5.6 mg/hm2 after biochar application, which proved the great potential for biochar in improving crop yield, especially to small farms with limited fertilizer availability. In addition to the effect of biochar itself, soil texture also has an important influence on crop yielded. Cluster analysis of biochar effects showed that the increase effect of biochar on crop yield occurred mostly in heavier texture soils rather than in lighter and medium texture soils. Jones et al. (2011) study showed that biochar had no effect on crop yield in the first year after application, but a significant yield increase was observed in the second and third years. Therefore, long-term studies on the application of biochar in soil are needed. Fertilizer plays an important role in agricultural harvest. After being applied to the soil, the fertilizer will be lost, which owns to natural reaction occurring in soil. Hence, the utilization of slow-releases fertilizers is an advantageous strategy to decline nutrient gas and leaching, especially large nutrient losses (N, P, K) (Wang et al., 2013). The conversion to biomass pyrolysis into biochar has an effective effect on reducing soil nutrient

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loss. The previous study found that biochar itself contained nutrients and helped promote plant growth. In most studies, when biochar is added to the soil, nutrients are released rapidly. If biochar absorbs external nutrients (N, P, K), it can be used as a slow-release fertilizer to provide nutrients (N, P, K). It is observed that biochar derived from lignocellulosic biomass contains lower plant nutrients, but it can be impregnated with external nutrients and then made into pellets. The final product can control the release of nutrients at a slower rate, thereby reducing nutrients loss (Kim et al., 2014). Phosphorus-containing biochar contains valuable nutrients, which can be used as a slow-release fertilizer to improve soil fertility and store C in the soil for a longer time. Mixing charcoal with ashes and immersing nutrients such as N, P, K in the wood residue can produce slow-release K and N fertilizers. Studies indicated that biochar slow-release fertilizers can be widely used in sustainable modern agriculture due to its effective nutrient retention characteristics. These biochar-based slow-release fertilizers, compound fertilizers, and granular slow-release fertilizers (N, P, K) still need to be fully evaluated. For example, field trials are extremely important before these materials are widely used in soil to support plant growth and development.

6. Conclusions and perspectives In summary, the global decline in soil and environmental quality has brought opportunities for biochar research. The research on the application of biochar in soil and agricultural environment has a very good potential for development. Its contribution to soil can not only effectively improve the environment, but also open up diversified modes of resource utilization, which is of great significance of soil treatment, maintenance of ecological balance and sustainable development of the agricultural environment. The current level of pollution in the agricultural environment has made the sustainable development strategy a great challenge, so biochar provides a new and effective solution for soil management. Numerous research results have shown that biochar has potential for agricultural soil improvement, carbon sequestration and emission reduction in agricultural fields, soil pollution prevention and control, and crop growth and yield to increase; it can play a positive role in improving agricultural soil and water environment, suppressing greenhouse effect, and contributing to sustainable ecological agriculture. However, there are still many shortcomings in the widespread application of biochar. In this paper, we propose a prospect of the following aspects: the response from different soil types, pollution status and fertility differences to the same biochar is very different, therefore, in the practical application process, suitable biochar should be prepared and screened according to the local conditions of different soils. Most of the current studies are centered on potted plants and indoor trials, and the effects of biochar on soil improvement, crops and the environment is a long-term process, and whether it is consistently effective against soil improvement and carbon sequestration and reduction needs to be further clarified by long-term positioned field trials. While applying biochar on a large scale to improve soil, crop yield, and environmental protection, it is also necessary to consider

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how to make charcoal production of low cost and good benefits. We should research and develop various special biochar-based fertilizers to realize the efficient use of biochar and provide strong support for the sustainable development of ecological agriculture. In view of the current status of biochar research, more attention should be paid to the following aspects in future research: First, the nature of biochar and its application effects to vary from different biomass materials and char conditions, and standardized and systematic research on biochar production should be actively carried out. Secondly, we should strengthen the research on long-term field positioning tests, conduct comprehensive long-term monitoring of the effects of biochar on different types of soil and crops, and conduct long-term verification of the long-term soil fertilization and crop yield to increase effects. The relevant research should not be limited to the summary of experimental laws, but should be considered comprehensively according to different soil conditions and climate types, and analyze the mechanism of the phenomenon formation in detail at the microscopic level. Thirdly, we should strengthen research on the microbial mechanism of biochar excitation on soil organic carbon and the interaction mechanism of biochar, organic carbon, and agglomerates. Fourth, as an excellent new slow-release material, biochar should be used as a basis for decentralized biochar production network, and research on the efficiency enhancement technology of biomass charcoal-based fertilizer and the effect of biochar and fertilizer dispensing should be accelerated to develop biomass charcoal-based fertilizer and soil conditioner to improve soil fertility and achieve sustainable development of agriculture and environment. With the widespread application of biochar, its negative impact on the environment should also arouse our attentions. The stability of biochar is one of the important characteristics of biochar applied to the environment. Since different biochar has unique physical and chemical properties, its stability still needs attention. The potential dissolution of organic matter in biochar under heavy metal complexation indicate that the existence of dissolved organic matter in biochar is caused by the instability of biochar. Biochar, especially sludge biochar, contains heavy metals, which will apply to the soil, causing heavy metal pollution. To promote the practical application of biochar, further research on its potential toxicity to the environment is necessary.

Acknowledgments The authors are grateful to the Shaanxi Introduced Talent Research Funding (A279021901 and F1020221012), China, Shaanxi Provincial Key R&D Plan Project (2022NY-052) and The Introduction of Talent Research Start-up fund (No. Z101021904), College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province 712100, China for the financial support. We are also thankful to all our laboratory colleagues and research staff members for their constructive advice and help.

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United Nations, 2013. World Population Prospects: The 2012 Revision. [WWW Document] Prospect 2012 Revises. Ussiri, D., Lal, R., 2013. The Role of Nitrous Oxide on Climate Change. Springer, Netherlands, pp. 1–28. Wainaina, S., Awasthi, M.K., Sarsaiya, S., Chen, H., Singh, E., Kumar, A., Ravindran, B., Awasthi, S.K., Liu, T., Duan, Y., Kumar, S., Zhang, Z., Taherzadeh, M.J., 2020. Resource recovery and circular economy from organic solid waste using aerobic and anaerobic digestion technologies. Bioresour. Technol. 301, 122778. Wang, S., Li, X.K., Lu, J., Li, H., Liu, B., Wu, Q., Wang, H., Xiao, G., Xue, X., Xu, Z., 2013. Effects of combined application of urea and controlled-release urea on yield, profifits of rapeseed and soil inorganic nitrogen. Chin. J. Oil Crop Sci. 35, 295–300. Wang, J., Liao, Z., Ifthikar, J., Shi, L., Du, Y., Zhu, J., Xi, S., Chen, Z., Chen, Z., 2017. Treatment of refractory contaminants by sludge-derived biochar/persulfate system via both adsorption and advanced oxidation process. Chemosphere 185, 754–763. Wang, C., Liu, J., Shen, J., Chen, D., Li, Y., Jiang, B., Wu, J., 2018. Effects of biochar amendment on net greenhouse gas emissions and soil fertility in a double rice cropping system: a 4-year field experiment. Agric. Ecosyst. Environ. 262, 83–96. Wang, D., Felice, M.L., Scow, K.M., 2020. Impacts and interactions of biochar and biosolids on agricultural soil microbial communities during dry and wet-dry cycles. Appl. Soil Ecol. 152, 103570. Wiedemeier, D.B., Abiven, S., Hockaday, W.C., Keiluweit, M., Kleber, M., Masiello, C.A., McBeath, A.V., Nico, P.S., Pyle, L.A., Schneider, M.P.W., Smernik, R.J., Wiesenberg, G.L.B., Schmidt, M.W.I., 2015. Aromaticity and degree of aromatic condensation of char. Org. Geochem. 78, 135–143. Wu, C., Cui, M.Q., Xue, S.G., 2018. Remediation of arsenic-contaminated paddy soil by iron-modified biochar. Environ. Sci. Pollut. Res. 25, 20792–20801. Xiao, R., Awasthi, M.K., Li, R., Park, J., Pensky, S.M., Wang, Q., Wang, J.J., Zhang, Z., 2017. Recent developments in biochar utilization as an additive in organic solid waste composting: a review. Bioresour. Technol. 246, 203–213. Xu, C., Liu, W., Sheng, G., 2008. Burned rice straw reduces the availability of clomazone to barnyardgrass. Sci. Total Environ. 392, 284–289. Xu, X., Cao, X., Zhao, L., 2013. Removal of Cu, Zn, and Cd from aqueous solutions by the dairy manurederived biochar. Environ. Sci. Pollut. Res. 20, 358–368. Xu, C., Xiang, Q., Zhu, H.H., 2018. Effect of biochar from peanut shell on speciation and availability of lead and zinc in an acidic paddy soil. Ecotoxicol. Environ. Saf. 164, 554–561. Xue, Y., Gao, B., Yao, Y., Inyang, M., Zhang, M., Zimmerman, A.R., Ro, K.S., 2012. Hydrogen peroxide modification enhances the ability of biochar (hydrochar) produced from hydrothermal carbonization of peanut hull to remove aqueous heavy metals: batch and column tests. Chem. Eng. J. 200-202, 673–680. Yang, Z.M., Fang, Z.Q., Zheng, L.C., 2016. Remediation of lead contaminated soil by biochar-supported nano-hydroxyapatite. Ecotoxicol. Environ. Saf. 132, 224–230. Yang, F., Gao, Y., Sun, L., 2018. Effective sorption of atrazine by biochar colloids and residues derived from different pyrolysis temperatures. Environ. Sci. Pollut. Res. 25, 18528–18539. Yousaf, B., Liu, G., Wang, R., Abbas, Q., Imtiaz, M., Liu, R., 2017. Investigating the biochar effects on C-mineralization and sequestration of carbon in soil compared with conventional amendments using the stable isotope (13C) approach. GCB Bioenergy 9, 1085–1099. Yu, X., Ying, G., Kookana, R.S., 2009. Reduced plant uptake of pesticides with biochar additions to soil. Chemosphere 76, 665–671. Yu, Z.H., Xie, L.K., Liu, S., 2014. Effects of biochar-manganese oxides composite on adsorption characteristics of Cu in red soil. Ecotoxicol. Environ. Saf. 23, 897–903.

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Yuan, C., Gao, B., Peng, Y., Gao, X., Fan, B., Chen, Q., 2021. A meta-analysis of heavy metal bioavailability response to biochar aging: importance of soil and biochar properties. Sci. Total Environ. 756, 144058. Yuvaraj, A., Thangaraj, R., Karmegam, N., Ravindran, B., Chang, S.W., Awasthi, M.K., Kannan, S., 2021. Activation of biochar through exoenzymes prompted by earthworms for vermibiochar production: a viable resource recovery option for heavy metal contaminated soils and water. Chemosphere 278, 130458. Zhang, M., Gao, B., Yao, Y., 2012. Synthesis, characterization, and environmental implications of graphenecoated biochar. Sci. Total Environ. 435-436, 567–572. Zhang, X., Fei, Y., Tian, X., Li, Y., 2016. The passivation effect of Pb, Cd composite polluted soil by phosphorus-modified biochar. Environ. Pollut. 39, 1017–1020. Zhang, K., Chen, L., Li, Y., 2017. The effects of combinations of biochar, lime, and organic fertilizer on nitrification and nitrifiers. Biol. Fertil. Soils 53, 77–87. Zhang, H., Shao, J., Zhang, S., 2020. Effect of phosphorus-modified biochars on immobilization of Cu (II), Cd (II), and As (V) in paddy soil. J. Hazard. Mater. 390, 121349. Zhang, Q., Duan, P., Gunina, A., Zhang, X., Yang, X., Kuzyakov, Y., Xiong, Z., 2021. Mitigation of carbon dioxide by accelerated sequestration from long-term biochar amended paddy soil. Soil Tillage Res. 209, 104955. Zheng, R., Chen, Z., Cai, C., 2015. Mitigating heavy metal accumulation into rice (Oryza sativa L.) using biochar amendment: a field experiment in Hunan, China. Environ. Sci. Pollut. Res. 22, 11097–11108. Zhou, N., Chen, H., Feng, Q., Yao, D., Chen, H., Wang, H., Zhou, Z., Li, H., Tian, Y., Lu, X., 2017. Effect of phosphoric acid on the surface properties and Pb(II) adsorption mechanisms of hydrochars prepared from fresh banana peels. J. Clean. Prod. 165, 221–230. Zhou, Y., Qin, S., Verma, S., Sar, T., Sarsaiya, S., Ravindran, B., Liu, T., Sindhu, R., Patel, A.K., Binod, P., Varjani, S., Singhnia, R.R., Zhang, Z., Awasthi, M.K., 2021. Production and beneficial impact of biochar for environmental application: a comprehensive review. Bioresour. Technol. 337, 125451. Zhu, Q., Kong, L., Shan, Y., Yao, X., Zhang, H., Xie, F., Ao, X., 2019. Effect of biochar on grain yield and leaf photosynthetic physiology of soybean cultivars with different phosphorus efficiencies. J. Integr. Agric. 18, 2242–2254.

12 Sustainable production and application of biochar for energy storage and conversion Quoc Cuong Doa,b, Thi Nhung Tranb, Thi Hien Tranc, D. Duong Lad, Huu Hao Ngoe, B.X. Thanhf, S. Woong Changg, and Dinh Duc Nguyeng CHEMICAL & PROCESS TECHNOLOGY DIVISION, K OR EA RESEARCH INSTITUTE O F CHEMI CAL TECHNOLOGY (K RICT), DAEJEON, REPUBLIC OF KOREA b DEPARTMENT OF CIVIL AND ENVIRONMENTAL ENGINEERING, KOREA ADV ANCED INSTITUTE OF SCIENCE AND TECHNOLOGY (KAIST), DAEJEON, REPUBLIC OF KOREA c INSTITUTE OF ENVIRONME NTAL SCIENCE, ENGINEERING AND MANAGEMENT, INDUSTRIAL UNIVERSITY O F HO CHI MINH CITY, HO CHI MINH CITY, VIETNAM d INSTITUTE OF C HEMISTRY AND MATERIALS, HANOI, VIETNAM e CENTRE FOR T ECHNOLOGY IN WATER AND WASTEWAT ER, SCHOOL OF CIVIL AND E N V I R ON M E NT A L EN G I NE E R I N G , U N I V E R SITY O F T ECHNOLOGY SYDNEY, SYDNEY, NSW, AUSTRAL IA f KEY LABORATORY O F ADVANCED WASTE TRE AT MENT TE CHNO LO GY , VI ET NAM NATIONAL UNIVERSITY HO C HI MINH (VNU-HC M) , HO CHI MINH CITY, V IETNAM g DEPARTMENT OF ENVIRONMENTAL E NERGY ENGINEERING, KYONGGI UNIVERSITY, SUWON, REP UB LIC OF KOREA a

1. Introduction Fossil fuels are among the most important energy sources today, accounting for more than 80% of the world’s primary total energy supply. However, owing to the rapid population growth and the fast industrial development, the demand for energy has escalated, which can lead to the depletion of fossil fuel resources in the future. In 2019, the total energy consumption of fossil energy sources worldwide was 583.9 EJ (oil: 193 EJ, gas: 141.5 EJ, and coal: 157.9 EJ) (Bernard, 2020). The growing global demand for fossil fuels, which simultaneously leads to the depletion of available fuel resources and negatively affects the environment (i.e., it causes the greenhouse effect and promotes climate change), poses great challenges for the energy industry. Therefore, many countries are studying alternative renewable energy sources and focus on the use of sustainable energy. In particular, renewable bioenergy sources are green energy sources, and carbon-containing materials for environment-friendly high-value energy storage have been successfully prepared from waste biomass. Scientists have transformed biomass into different states (solid, liquid, or gaseous) (Oncioiu et al., 2020). Biomass has been effectively reused in many different ways: as green renewable energy source (Perlack et al., 2005), biochar/ hydrochar (Rangabhashiyam and Balasubramanian, 2019; Xiao et al., 2012), adsorbent Current Developments in Biotechnology and Bioengineering. https://doi.org/10.1016/B978-0-323-91873-2.00011-X Copyright © 2023 Elsevier Inc. All rights reserved.

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materials (Tran et al., 2020, 2021), porous carbon materials (Isahak et al., 2013; NavarroSua´rez et al., 2014), catalyst materials (de Almeida Ribeiro et al., 2020), and energy storage materials. The term “biochar/hydrochar” has coemerged with the terms “renewable energy,” “soil conditioner,” and “carbon adsorbent” (Singh et al., 2021). The International Biochar Initiative defined biochar as “a solid material obtained from the thermochemical conversion of biomass in an oxygen-limited environment” (IBI, 2013). Some scientists base their definitions of materials on the composition, source of raw materials, method of preparation, and application (Nartey and Zhao, 2014); for example, the term “charcoal” represents a product created through partial combustion of biomass (plants or trees) (Mulabagal et al., 2020). Therefore, it is important to differentiate the terms “biochar,” “charcoal,” and “hydrochar” based on their production processes. “Charcoal” is a carbon-rich solid product created through the combustion of biomass and used as a fuel source for energy production, and “biochar” is an alternative term for “charcoal” when used for a specific purpose such as a soil conditioner and valuable carbon material (Panwar et al., 2019). Biochar is dry and produced via pyrolysis or microwave treatments, and hydrochar is a sludgy mixture of solid and liquid char produced via hydrothermal carbonization (HTC) (Becker et al., 2014; Erdogan et al., 2015; Kambo and €kela € et al., 2015; Nizamuddin et al., 2016; Oliveira et al., Dutta, 2015; Kang et al., 2012; Ma 2013; Wei et al., 2011; Wu et al., 2017). Hydrochar is similar to biochar; however, their physicochemical properties significantly differ (Al-Wabel et al., 2019). Both materials have originally been developed and used as effective soil conditioners (Li et al., 2019). However, the great recent progress in research and development of technology for pyrolysis and HTC have extended the application of biochar and hydrochar (Kambo and Dutta, 2015). Their common application fields include water purification (Wang et al., 2020a, c), fertilizers (Carneiro et al., 2021), sorption (Tran et al., 2020, 2021), energy storage and conversion (Liu et al., 2019), and catalysis (Gholizadeh et al., 2021). Therefore, biochar and hydrochar are considered the simplest solution for the development of sustainable energy sources and environmental management (Maniscalco et al., 2020). The use of biochar and hydrochar carbon precursors from biomass as gas storage, energy storage, and conversion materials involves many challenges. Human activities result in the emission of substantial amounts of toxic gasses (e.g., CO2, CH4, N2O, and SO2), which are one of the main causes of global climate change. The prevention of their emission has been studied with different adsorption methods, such as pressure swing adsorption (Bernardo et al., 2021), vacuum pressure swing adsorption (Majchrzak-Kucęba and Sołtysik, 2020), temperature swing adsorption, and vacuum swing adsorption (Ghanbari and Kamath, 2019). Solid adsorbents derived from biochar/hydrochar are among the most promising alternatives (Maniscalco et al., 2020). They must absorb large gas amounts, have high absorption rates, be easily available and regenerable, and exhibit good selectivity and low cost (Gonza´lez et al., 2013). However, the production of biochar/hydrochar-derived sorbents that meet all these requirements is challenging. Some scientists have proposed the production of biochar/hydrochar-derived porous carbon from renewable waste biomass sources as adsorbent to capture CO2, hydrogen, and CH4; this method is considered very

Chapter 12 • Sustainable production and application of biochar 335

promising. The researchers (Acevedo et al., 2020; Bernardo et al., 2021; Jain et al., 2016) focused on the CO2 adsorption capacity of different biochar/hydrochar-derived carbon materials. In addition, different carbon materials such as carbon aerogels (Zhong et al., 2018), graphene (Ahamed et al., 2021), and carbon nanotubes (Orinˇa´kova´ and Orinˇa´k, 2011) have been used to adsorb and store hydrogen owing to their microporous structures. Biochar/hydrochar-derived porous carbon from low-cost renewable biomass sources is a good alternative for storing harmful gasses that negatively affect the global climate. Furthermore, biochar/hydrochar-derived carbon from renewable biomass sources can be used for the production of energy storage materials such as batteries and supercapacitors. For example, hydrochar obtained via hydrothermal carbonation can be used as electrode material in Li- and Na-ion batteries (Fakkaew, 2016; Maniscalco et al., 2020). For the latter, there are many types of electrodes made from hydrochar-derived materials such as anodes based on intercalation/deintercalation (Park et al., 2011; Tang et al., 2012), anodes based on alloying reactions (Lou et al., 2009; Yu et al., 2011), anodes based on conversion reactions ( Jiang et al., 2011; Qi et al., 2011), and positive electrodes (Ellis et al., 2010). Lithium batteries have revolutionized the mobile electronics industry owing to their highly efficient recharging capabilities (Liang et al., 2019). Compared to other conventional batteries with identical volumes and weights, lithium batteries have 2–3 times more energy capacity. In addition, many researchers have investigated the use of hydrochar-derived materials as electrodes in Na-ion batteries (Deng, 2015). Chevrier and Ceder (2011) compared the characteristics of Na- and Li-ion batteries (LIBs); accordingly, Na-ion batteries have the voltage, stability, and diffusion barrier characteristics of Na-ionic materials, which make them as good as Li-ion batteries (Palomares et al., 2012). In addition, Na-ion batteries contain latent energy that can promote the use of renewable energy sources as the main source of energy rather than a supplement for low-cost energy storage systems (Palomares et al., 2012). The increasingly advanced society and the introduction of electric vehicles have been an important turning point for Li-S batteries (Azimi et al., 2015) and supercapacitors (Lin et al., 2021) because Li-ion batteries do not meet the requirements for long-term energy storage. Biomass-derived materials are used as electrodes in Li-S batteries (Brun et al., 2013; Zhi et al., 2008); when a complete reaction is assumed in Li-S cells, they can provide significantly higher energy densities (2500 Wh kg 1 or 2800 Wh L 1) per unit of weight or volume. However, Li-S batteries have deficiencies that prevent their widespread application in practice (Robinson et al., 2021). Electrochemical capacitors and supercapacitors have attracted much attention in research owing to their good electrochemical energy storage characteristics, high energy densities, long lifespans, and wide operating temperature ranges (Lin et al., 2021). Electrochemical capacitors can be classified into two types of double-layer capacitors (EDLC), and dummy capacitors depending on their charge storage mechanism (Conway, 2013). EDLCs provide high capacitance values owing to their large surface areas; they are the most common form of EDLCs that can be made from biochar/hydrocharderived porous carbon materials (Yang et al., 2019). Biochar/hydrochar-derived porous carbon materials from renewable biomass sources have large specific surface areas and

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can be easily activated; in addition, adding surface functional groups to the surface can increase the capacitance value (Yang et al., 2019). Therefore, biochar/hydrochar-derived porous carbon materials from renewable biomass sources are the top candidates for the production of supercapacitors with outstanding characteristics such as high flexibility regarding porosity and large surface areas, suitable chemical characteristics of the surface, good conductivity, eco-friendliness, low cost, and abundance. To minimize the factors affecting climate change, scientists have investigated the sustainable use of biochar/ hydrochar for energy storage applications and the conversion of renewable sources.

2. Biochar production technology 2.1 Thermochemical processes Biochar can be produced via a thermochemical conversion process by heating biomass or wastes containing carbon at high temperature (150–1100°C) in a deficit or an oxygen-free environment. The characteristics of biomass feedstock and the thermochemical operating parameters (i.e., the temperature range, heating rate, vapor residence time, and use of oxygen) affect the yield and physicochemical properties of the resulting biochar. In general, biochar prepared from agricultural residues provides a higher yield and carbon content than that derived from wood-based materials (Wang et al., 2013). Animal manure, food waste, and paper mill waste can yield higher biochar yields with higher ash and volatile carbon contents than crop residues and wood biomass (Enders et al., 2012). In addition, the higher lignin content in biomass results in a higher yield of biochar with a higher carbon content because of its weak thermal decomposition (Qu et al., 2011; Wang et al., 2013). The higher thermal treatment temperature can increase the surface area, alkalinity, pH value, ash content, and aromaticity and decrease the yield, number of surface functional groups, and cation exchange capacity of the resulting biochar (Wiedemeier et al., 2015; Wu et al., 2019; Zhang et al., 2015). The thermal treatment of biomass at a moderate temperature (up to 500°C) for a long vapor residence time (up to 4 h) can result in biochar and bio-oil (typically 15%–35% and 30%–50% biomass feedstock is used, respectively); in addition, higher bio-oil yields (50%–70%) can be obtained with a shorter residence time (up to 2 s) of the vapor in the reactor (Panwar et al., 2019). The energy requirements for the production of biochar from biomass ranges from 1.1 to 16 MJ/kg biochar (Panwar et al., 2019). Ro et al. (2010) discovered that a huge amount of energy is needed to dry animal manure before pyrolysis; more specifically, 232.3 MJ is needed to produce 1 kg biochar from wet swine manure with 97% moisture content; the required energy can be significantly reduced when the manure is dewatered to a 75% moisture content (12.5 MJ/kg biochar). Since the proper classification is not available in literature, the thermochemical processes of interest in this chapter are limited to biochar production: pyrolysis, torrefaction, carbonization, and gasification processes. Each process produces different quantities of biochar products with different qualities. The total solid carbon amount in biochar obtained from biomass ranges between 5% for gasification and approximately 90% for torrefaction (McHenry, 2014). The main differences between these techniques are highlighted in Table 1.

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Table 1 Product distribution of biomass conversion in different thermochemical processes. Product distribution (wt.%)

Process

Temperature (°C)

Heating rate (°C/min)

Residence time

Solid (biochar)

Liquid biooil)

Gas (syngas)

Slow pyrolysis

300–700

10

Minutes to hours

15–40

30–50

30–35

Fast pyrolysis

300–1000

>100

0.5–2 s

12–25

50–75

13–25

Dry torrefaction Wet torrefaction

200–350

10

70–90

5–20

5–10

175–260

10

Minutes to hours Minutes to hours

70–90

5–20

0–5

Gasification

700–900

>100

Seconds to minutes

5–10

5–10

85

Hydrothermal carbonization Flash carbonization

150–400

10

Hours

50–80

5–40

2–10

300–600

10

100°C/min) and a short vapor residence time (50°C/min) and with 30–60 min residence time; WT is conducted in pressurized water or diluted acid solutions at up to 5 MPa and 180–260°C with 5 min to several hours of residence time (Chen et al., 2021; Yan et al., 2010); DT is more suitable for biomass feedstocks with low moisture contents; therefore, a biomass predrying step is required. By contrast, WT is preferred for the treatment of biomass with high moisture contents (e.g., urban waste, animal manures, and sewage sludge); the torrefied product must be dried in a drying process. In DT, the biomass color changes to dark brown. The three primary components of biomass (i.e., cellulose, hemicellulose, and lignin) decompose with different kinetics (Sarvaramini et al., 2013). Moreover, the property of biomass changes from hydrophilic to hydrophobic owing to the degradation of hemicellulose. Moisture dissipation, functional groups, and carbon dioxide formation are caused by the polymerization of cellulose (Acharya et al., 2015). This process can yield a maximal mass yield of 70% and a maximal energy content of 90% of those of the original biomass (Acharya et al., 2015; Mamvura and Danha, 2020). However, the highly volatile compounds and low moisture content of this biochar type increase the risk of spontaneous ignition and, therefore, fire or explosion (Chen et al., 2018). In addition, the high ash content in the biochar can limit its application range (Chen et al., 2018). In WT, the main factors affecting the yield and characteristics of the WT product are the operation temperature, pressure, residence time, atmosphere, and liquid medium (Chen et al., 2018). Researchers have studied the WT of lignocellulosic biomass (Yan et al., 2009, 2010) and discovered that the hydrolysis of cellulose and hemicellulose is the main reaction occurring in the system; it strongly depends on the operation temperature rather than the reaction time, moisture content, and particle size of biomass. WT can provide a solid product with a higher yield and energy density than those obtained with DT: 88.3% of the mass and 89.1% of the energy of the biomass feedstock can be retained (Chen et al., 2018; Yan et al., 2009).

2.1.3 Carbonization Hydrothermal carbonization (HTC; i.e., wet pyrolysis) is a thermal treatment process designed for the conversion of wet biomass feedstocks into highly dense carbonaceous residue. In HTC, organic substances are processed at autogenous pressure in a closed aqueous environment at temperatures between 150 and 350°C to form condensed carbonaceous solid products (i.e., hydrochar) and a liquid stream rich in organic compounds (Masˇek, 2016). Biochar derived via HTC consists mainly of alkane structures with low stability (Chen et al., 2019; Weber and Quicker, 2018). The benefits of HTC over pyrolysis or torrefaction include the higher C/O ratio and energy density, better grindability, and higher degree of hydrophobicity of the obtained biochar (Gabhane et al., 2020).

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In flash carbonization (thermochemical process), biomass is quickly and efficiently converted into biochar. The process starts with the ignition of a flash fire at the bottom of a packed biomass bed at elevated pressure ( 2 MPa) at 300–600°C for a residence time of less than 30 min (Chen et al., 2019; Wade et al., 2006). There are only few reports on flash carbonization techniques in literature, and the method is not commonly applied for biochar production (Yaashikaa et al., 2020). According to a typical experiment on flash carbonization of waste oak wood floorboards presented by Antal et al. (2003), the solid product can retain 40% of the mass and 66% of the energy content of the biomass feedstock.

2.1.4 Gasification In gasification (thermochemical process), biomass materials are converted into a gaseous mixture (i.e., synthesis gas or syngas) at relatively high temperature (700–900°C) in a lowoxygen concentration environment. Most gasification systems exhibit very high carbon conversion efficiencies (94%–99%) and retain only a relatively small solid portion as a byproduct with a low carbon content; thus, this method is less suitable for biochar production than pyrolysis (Masˇek, 2016). Nevertheless, Pels et al. (2005) reported that up to 70% carbon content can be achieved in fly-ash through fluidized bed gasification. However, the material can contain high concentrations of harmful organic and inorganic contaminants (Masˇek, 2016). Biochar produced via gasification generally has higher contents of aromatic compounds, higher degrees of porosity, and larger specific surface areas than biochar derived via pyrolysis (Meyer et al., 2011; Zhang et al., 2019c).

2.2 Operation mode Biochar-producing reactors can be operated in the batch, semibatch, or continuous mode.

2.2.1 Batch mode The batch process is a well-known technique that is commonly used in small-scale biochar production reactors. This process starts with a heating period in which the biochar is produced, followed by a nonproductive cooling period for biochar discharge and the preparation of the next batch. Heat recovery is not performed in these reactors, and the produced biochar can only be discharged after it has cooled down to a suitable temperature. Biochar production via the batch mode can retain 30% of the biomass feedstock. Despite the low construction costs of batch-mode reactors, the operation costs for the repeated heating and reheating processes of the oven are high. Moreover, the volatile products generated during this process are usually emitted into the atmosphere, thereby potentially affecting human health and the environment. A typical batch process includes the use of earthen mound kilns, concrete and metal kilns, bricks, and retorts (Panwar et al., 2019).

2.2.2 Semibatch mode The semibatch configuration involves a series of batch reactors that improve the performance of the heating oven (de Almeida Ribeiro et al., 2020). The heat-containing vapor

Chapter 12 • Sustainable production and application of biochar 341

stream in the system is recycled in between the successively installed batches. The production capacity of a semibatch system depends on the number of integrated batch runs. A typical example is the semibatch system of the Carbo twin retort, which was initially developed in the Netherlands in the 1990s; it consists of two retorts in an insulated oven (Garcia-Nunez et al., 2017). The systems have low labor requirements, high energy efficiency, and high yields; in addition, the quality of the biochar is high, the system can be easily scaled-up, and it emits low emissions.

2.2.3 Continuous mode The continuous operating mode is designed to ensure the continuous flow of biomass in the reactor with an axial temperature profile; the material undergoes entire subprocesses, such as drying, preheating, pyrolysis, cooling, and discharging steps. This technology is commonly used in commercial-scale production owing to its high productivity, energy efficiency, and high quality of the resulting biochar (Duku et al., 2011; Panwar et al., 2019). These systems offer great benefits regarding the flexibility of the use of biomass feedstocks and large-scale production (Gwenzi et al., 2015). Typical continuous pyrolysis reactors are the drum-type pyrolyzer, screw-type pyrolyzer, and rotary kiln.

2.3 Biochar activation methods Biochar derived from biomass pyrolysis generally has a small surface area, poor porosity, and few surface functional groups. Particularly, volatile organic compounds can be retained in the solid product obtained in a low-temperature conversion process (e.g., torrefaction); the resulting properties are similar to those of a product between biochar and raw biomass. However, the surface functional groups can be removed at high temperature, which may limit the application of the resulting biochar. Consequently, researchers have investigated solutions for producing biochar with desired properties to increase its applicability. Many activation techniques have been applied to improve the functionality of biochar.

2.3.1 Physical activation Physical or thermal activation aims to modify the physical structure of raw biochar by partially gasifying the feedstock at high temperature (e.g., 700–900°C) in an oxidizing atmosphere such as air, ozone, carbon dioxide, and steam (Prauchner and Rodrı´guez-Reinoso, 2012; Zhang et al., 2017a, b). The oxidizing agents can penetrate the internal structure of raw biochar, thereby gasifying the volatile carbon content. The elimination of carbon atoms can lead to the opening of closed pores and the generation of new pores, thereby resulting in a larger specific surface area, a more robust microporous structure, and abundant functional groups in the obtained activated biochar. The properties of activated biochar highly depend on the activation temperature, duration, type of biomass precursor, and activation agent. In general, a higher degree of porosity and a larger pore size

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distribution can be obtained at higher activation temperatures and with longer durations. The specific surface area and pore volume significantly increase when biochar has been activated in steam or a carbon dioxide atmosphere (Chen et al., 2020; Kołtowski et al., 2017; Lima et al., 2010), whereas a high ash content and a low product yield are obtained when the activation has been done in air (Sakhiya et al., 2020). Physical activation methods are simple but energy-intensive because of the high operating temperatures. Therefore, many different strategies have been proposed to improve the physicochemical properties of biochar at a reasonable cost; the methods include microwave activation, ultrasound irradiation, plasma treatments, and electrochemical modification (Sajjadi et al., 2018).

2.3.2 Chemical activation Chemical activation aims to modify biochar functionalities when it involves a thermal treatment process with a chemical activating agent (e.g., acid, base, H2O2, and KMnO4) with a specific duration. This process significantly affects the surface chemistry of biochar depending on the type of activating agent, temperature, and activation time. The chemical activation methods of biochar can be further classified into oxidation, amination, sulfonation, metal oxide impregnation, and magnetization (Sajjadi et al., 2019). Surface oxidation aims to increase the number of oxygen-containing functional groups on the surface of biochar, which can be accomplished via activation with oxidizing gasses or acidic/basic oxidizing solutions. Surface amination of biochar is usually performed in the presence of ammonia or amino-containing reagents to create basic amino moieties on the surface. Surface sulfonation is used to introduce sulfonic groups (SO3H) in biochar by fuming it or treating it with concentrated sulfuric acid. Metal/metal oxide impregnated biochar can be created with various preparation techniques (e.g., wet impregnation, depositionprecipitation, and hydrothermal methods) and a solution containing targeted metal salt. Powdered biochar can be magnetized to recover it from environmental media. This is usually accomplished by integrating magnetic nanoparticles into biochar structures, which can be easily recycled with an external magnetic field. Chemical activation usually presents several advantages over physical activation, such as the lower temperature, higher yield, and larger surface area of the obtained biochar (Sakhiya et al., 2020). However, the harmful effects of the used chemical agents must be monitored during the reaction.

2.3.3 Biological activation Biochar can be activated with microorganisms through colonization and biofilm generation on its porous structures; thus, biochar with desired properties can be engineered. The microorganisms can penetrate the small pores of biochar and develop a biofilm; the film structure prevents them from being washed out (Kazemi Shariat Panahi et al., 2020). Biologically activated biochar is usually applied for environmental remediation purposes; the material can be used as an adsorbent, a catalyst, and an ion exchanger (Dalahmeh et al., 2018; Frankel et al., 2016; Yao et al., 2018).

Chapter 12 • Sustainable production and application of biochar 343

3. Biochar market Biochar is commercially available and has been utilized for a long time. Numerous types of biomass (including agriculture and forestry by-products, industrial by-products, animal manure, and sewage sludge) have been successfully employed for biochar production. The commercial biochar production and distribution is growing sustainably worldwide, thereby opening up more opportunities for new business models. The total volume of biochar traded in 2015 was 85,000 tons according to a survey comprising 326 companies (“State of the biochar industry—2015,” International Biochar Initiative, 2015). The global biochar market size is expected to increase to USD 3.82 billion in 2025 (“Biochar Market by Feedstock Type (Woody Biomass, Agricultural Waste, Animal Manure, and Others), by Technology (Pyrolysis, Gasification, and Others), and by Application (Electricity Generation, Agriculture, and Forestry): Global Industry Perspective, Comprehensive Analysis, and Forecast, 2018–2025” Zion Market Research, 2019). The demand for biochar is growing because of its increased usage in many application sectors including power generation, energy storage, agriculture, forestry, and environmental treatment techniques. The increase in the organic food demand is expected to be the main driver of the market during this period. In addition, many other important factors such as positive government policies, affordable biomass feedstocks, and environmental awareness of waste managers contribute to the success of the worldwide biochar industry. The market price of biochar is in the range of 90–8850 USD/ton (Ahmed et al., 2016; Thompson et al., 2016); it depends on the selected biomass feedstock, manufacturing conditions, transportation conditions, and application of biochar. Homagain et al. (2016) conducted a lifecycle cost and economic assessment analysis of biochar-based bioenergy production technology in Northwestern Ontario (Canada); they revealed that the pyrolysis process is the most significant cost factor; it accounts for 36% of the total cost of the system; land application, feedstock collection, and transportation costs account for 14%, 12%, and 9% of the total production cost, respectively. North America, where most biochar types are made from residues of the wood and energy industries (Miles, 2020), was estimated to be the biggest biochar market in 2018, followed by Europe and Asia-Pacific (“Biochar Market by Feedstock Type (Woody Biomass, Agricultural Waste, Animal Manure, and Others), by Technology (Pyrolysis, Gasification, and Others), and by Application (Electricity Generation, Agriculture, and Forestry): Global Industry Perspective, Comprehensive Analysis, and Forecast, 2018–2025,” Zion Market Research, 2019). By contrast, the majority of biochar production systems in many developing countries are usually micro-scale (household biochar cook stoves) to small-scale (village-level) systems, which have received very little attention in economic analysis. Most companies that are listed online and commercially trade biochar are located in developed countries (i.e., the United States, Canada, Australia, and several Western European countries) and a few developing countries (i.e., India and Turkey) (Alhashimi and Aktas, 2017).

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4. Applications of biochar in energy storage and conversion fields 4.1 Application of biochar in supercapacitors Supercapacitors are promising electrochemical energy storage systems with high energy and power densities, long lifecycles, and excellent reversibility characteristics; they are widely applied in automobiles, wireless terminals, power supplies, and power grids. However, supercapacitors are not completely renewable and sustainable since their core elements (electrodes) are composed of fossil fuel-derived carbon materials (e.g., carbon blacks, graphite, carbon nanotubes, and conductive polymers) and inorganic materials containing rare metals (Kaushal et al., 2018; Lee et al., 2018). Therefore, the development of alternative electrodes based on biomass-derived materials is considered a potential solution for improving economic and environmental factors. The performance characteristics of supercapacitors depend on the electrical conductivity, porosity, surface area, and surface oxygen-rich functional groups of the electrode material (Hassan et al., 2020; Kumar et al., 2020). In general, electrode materials with high electronic conductivity values, good chemical stability, large surface areas, and abundant pores and oxygencontaining functional groups are suitable candidates for the preparation of highcapacitance and high-energy density supercapacitors. Biochar-based materials derived from different biomass types, such as bamboo (Gong et al., 2017), coconut shells (Sun et al., 2013; Xia et al., 2018), pomelo peals (Wang et al., 2018a, b, c), rice straws (Ding et al., 2020; Liu et al., 2018), water hyacinths (Zheng et al., 2017), softwood kraft pulp (Abouelamaiem et al., 2018), durian shells (Wang et al., 2020a, c), coffee grounds (Choi et al., 2018), and tobacco rods (Zhao et al., 2016), have been developed through thermochemical and/or chemical activation processes for the preparation of supercapacitor electrodes. The activation of biochar is crucial for attaining devices with high capacitance values. Gupta et al. (2015) proposed a low-temperature (97% naphthalene and >98% fluorene are present

Amstaetter et al. (2012)

benzo(b) fluoranthene benzo(a) pyrene anthracene benzo(a)anthracene benzo(g,h,i)perylene benzo(k)fluoranthene Chrysene Fluoranthene Indeno (1,2,3-cd)pyrene Phanenthrene Pyrene

Coconut shellsbased PAC and anthracite-based powder activated carbon (PAC)

5 g sediment containing 30  6 mg/kg total PAHs (pyrene 3.8 mg/kg) was mixed with 50 mg activated carbon (AC) in a 40 mL solution and shaken for 30 d

Naphthalene, fluorene

Activated carbon (AC)

2–5-ring PAHs

Biochar

PAHs (polycyclic aromatic hydrocarbons) are a group of chemicals (excluding naphthalene among 16 USEPA PAHs) Naphthalene

PAC/GAC

At 25°C for 15 min, 12 mg/L naphthalene and 1 mg/L of fluorene were combined with 50 mg CO2-activated petroleum cokes 5 g of polluted soil in 10 mL of acetone and hexane in a 1:1 ratio for extraction, the solution was shaken at 200 rpm for 2 h 2% AC was added to urban soil with a concentration of 38 mg/kg of freely dissolved S15 PAH 15, 9, 7.5, 7.5, 2.5, 3, 4, and 4 mg of biochar from at 100, 200, 250, 300, 400, 500, 600, and 700° C, respectively, having a surface area of 0.65–490.8 m2/g mixed with h initial naphthalene concentrations of 0.03–0.94 and agitated at 20 rpm for 3 d PAHt0 0.5 mg/kg; biochar dosage 0.1–0.5 (w/w%); remediation time 4 d

Phenanthrene

Pine needle biochar

Biochar

Awoyemi (2011)

After 60 days, >50% of HMW PAHs and >40% of LMW PAHs decreased

Beesley et al. (2010)

PAC removed 99% of aqueous PAHs, while GAC removed 64%

Bra€ndli et al. (2008)

Maximum adsorption is 136.8 mg/g

Chen et al. (2008)

11–60

Cui et al. (2011)

Continued

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Current Developments in Biotechnology and Bioengineering

Table 2

Removal of polyaromatic hydrocarbons by biochar—cont’d

Types of PAHs

Biochar/ Material

P16

Biochar

Acenaphthylene

Biochar

Anthracene benzo(a)pyrene benzo(a)A benzo(b)fluoranthene benzo(k)fluoranthen Chrysene, fluorene Ondeno (1,2,3-cd) pyrene benzo(g,h,i)perylene Phenanthrene, pyrene Anthracene benzo[a]anthracene Fluoranthene Phenanthrene Chrysene Fluorene Pyrene Phenanthrene; pyrene

PAC/GAC

Parameters

PAHs removal (%)

References

PAHt0 8.0–15.4 mg/kg; biochar dosage 5.0 (w/w%); remediation time 28 d Biochar dosage 5.0 (w/w %); remediation time 28 d 2% by weight AC was added to the soil where pore water PAHs concentration was 31 ng/L

98

GomezEyles and Ghosh (2018) GomezEyles et al. (2013) Hale et al. (2012)

80

After 17 months, PAC and GAC scored 93% and 84%, respectively, and after 28 months, PAC and GAC scored 76% and 69%

Biochar

PAHt0 1458.0 mg/kg; biochar dosage 15.0 (w/ w%); remediation time 1d

>99

Ho et al. (2004)

Biochar

Biochar dosage 1 (w/w%); remediation time 56 d Phenanthrene 9.07 mg/L, fluorene 10.05 mg/L, pyrene 10.57 mg/L, Triton X-100 4.55 g/L, and 1e8 g/L biochar shaped at 400, 600, and 800°C, respectively, with surface areas of 427, 537, and 652 m2/g At 700°C, maximum adsorption occurs on Biochar. 0.5%–10% AC and biochar combined with sewage sludge containing 13.2 ng/L PAHs dissolved

52–69

Jia et al. (2020)

95.8%–98.6% by biochar 800°C, 71.8%– 88.1% for biochar 400°C and 82.4%–93.4% for biochar 600°C and Triton X-100 recovery >87%

Li et al. (2014)

A.C. removes 56%–95% of contaminants, while biochar removes 0%– 57%.

Oleszczuk et al. (2012)

Phenanthrene Fluoranthene Pyrene

Biochar

Anthracene, B(g,h,i) Perylene benzo(a)pyrene benzo(a)anthracene, Chrysene benzo(b)fluoranthene benzo(k)Fluoranthene Indeno(1,2,3 cd)pyrene Naphthalene Phenanthrene Pyrene

AC/Biochar

Chapter 13 • Role of biochar in polyaromatic hydrocarbons remediation

Table 2

373

Removal of polyaromatic hydrocarbons by biochar—cont’d

Types of PAHs

Biochar/ Material

Anthracene Phenanthrene; pyrene

Biochar

Naphthalene, phenanthrene, pyrene

Rice husk activated carbon (RHAC)

Pyrene

Biochar

Parameters

PAHs removal (%)

References

PAHt0 10.0 mg/kg; biochar dosage 4.0 (w/w%); remediation time 30 d 5 mL of 8 mg/L of each PAH was combined with 2 mg of activated RHAC and shaken at 200 rpm for 1–7 d for the kinetic test. RHAC 0.1–7 mg was combined with 5 mL of 8 mg/L per PAH and shaken at 25  1°C until equilibrium was reached PAHt0 < 0.1 mg/kg; biochar dosage 1.3 (w/w%); remediation time 100 d

>95

Silvani et al. (2017)

Maximum pyrene adsorption is 104.5 mg/g

Yakout et al. (2013)

29%

Chen et al. (2019a, b)

relative solubility) between suspension and water by a factor of two (Cui et al., 2011; Marchal et al., 2013). PAH sorption of about 80% was observed after 28 days of operation in canal sediments when added 5% peanut hull biochar (Gomez-Eyles et al., 2013). About 98% PAH reduction was attained in pore water using 5% pine wood biochar after 28 days in marine sediments (Gomez-Eyles and Ghosh, 2018). Sorption of PAH was increased by 99% after adding 15% coconut biochar to marine sediments (Ho and McKay, 1998). Similar effects of biochar doses on sorption capacity were reported in other studies as well (Chen et al., 2019a, b; Silvani et al., 2017), respectively, used 1.25% of a macadamia nutshell, 1% of mangrove plant and 4% of pinewood biochar for the sorption of pyridine (PYR), phenenthene (PHE) + PYR + and anthracene (ANT) + PYR + PHE. When PYR was individually treated, 29% sorption was observed after 100 days (Chen et al., 2019a, b), whereas sorption increased to 52% for PYR + PHE mixture after 56 days ( Jia et al., 2020), and for PYR + ANT + PHE, 95% sorption was achieved after 30 days (Silvani et al., 2017). The results show that PAHs (such as PYR) show better sorption when present in the mixture, which is likely due to their reduced solubility due to increased quantity. Reportedly, the high salinity of marine sediments, high specific surface area (343 m2 g L 1), biochar dose (4%) and initial concentration of PAH improve the sorption mechanism (Silvani et al., 2017). Adsorption kinetics: Designing an effective sorption system for any recalcitrant compound such as PAHs requires the adsorption rate, which could be determined through kinetic studies by assessing the minimum time needed to attain equilibrium adsorption. Among the various models studied for PAHs’ adsorption kinetics, the most common are pseudo-first-order kinetics, pseudo-second-order kinetics Elovich equation. In most

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Current Developments in Biotechnology and Bioengineering

cases, the pseudo-first order equation is generally evident only during the initial 20–30 min of the sorption process and not for the entire reaction time (Ho and McKay, 1998). In chemisorption processes, the pseudo-second-order equation might be well applicable with better correlation (Ho and McKay, 1998). It has been reported in the literature that for PAH adsorption onto various adsorbents (zeolite, activated carbon, organo-sepiolite, NH2-SBA-15 and plant residue-based sorbent), experimental data fits appropriately with the pseudo-second-order kinetic model (Balati et al., 2015; Cabal et al., 2009; Chang et al., 2004). Besides pseudo-second-order kinetics, PAHs’ adsorption such as acenaphthene, phenanthrene, anthracene, fluoranthene onto soil has been known to follow the Elovich equation (Rivas et al., 2008). It was observed that the sorption rate was high at the beginning, which subsequently slowed down at last.

3.2 Role of biochar addition in availability of polyaromatic hydrocarbons The portion or aspect of PAHs that is accessible to organisms is referred to as bioavailability. Bioavailability is an essential parameter for deciding PAHs’degradation fate in the environment (Ortega-Calvo et al., 2013). Passive samplers (e.g., polythene, silicone), terrestrial invertebrates (e.g., Chironomus plumosus) and mild solvents (e.g., butanol, Tenax) can all be used to detect PAH in polluted sites as dissolved, biodegraded, and bioaccumulated fractions (Yang et al., 2016). The addition of biochar to sediments reduces PAHs’ bioavailability, indicating reduced environmental risk linked with PAHs presence (Cui et al., 2013). The bioavailability of PHE in river sediments decreased by 72% on the addition of 1.0% corn stalk biochar (Wang et al., 2011), while 5% wood biochar decreased PAH bioavailability in sediments with a lower sediment accumulation factor as compared to control (biochar free) (Shen et al., 2012). Deviating from the results of Wang et al. (2011) and Shen et al. (2012), the need for more biochar (>5%) for a similar reduction in PAH bioavailability was reported by Gomez-Eyles et al. (2013) and Gomez-Eyles and Ghosh (2018) through the addition of activated carbon. This difference in results may be attributable to higher biochar pore availability (Silvani et al., 2017), resulting in better PAH sorption (Verheijen et al., 2010), while the higher biochar percentage used by Gomez-Eyles et al. (2013) physically limited PAH transfer to biochar (Oleszczuk et al., 2012). In a study by Han et al. (2015), the sorption performance of conventionally pyrolyzed biochar was compared with magnetized biochar in river sediments. A 50% reduction in PAH bioavailability was observed with conventional biochar, whereas no considerable reduction was noted with magnetized biochar after 30 days of operation. Inefficient output with magnetized biochar was most likely due to pore obstruction by magnetite and reduced intraparticle diffusion (Han et al., 2015). When the biochar sediment mixture was shaken at 100 rpm, however, the interaction of magnetized biochar with PAHs increased, resulting in a 68% reduction in PAH bioavailability compared to the control samples (Han et al., 2017). The effects of wheat straw biochar (3%) generated at 400 and 700°C for 4 h on PAH bioavailability in river sediments were investigated (Chi and Liu, 2016). After 54 days, biochar obtained at 700°C reduced PHE bioavailability by 69% and PYR bioavailability by 55%, while

Chapter 13 • Role of biochar in polyaromatic hydrocarbons remediation

375

biochar obtained at 400°C reduced PYR and PHE bioavailability by 29% and 23%, respectively (Chi and Liu, 2016). The bioavailability of PYR and PHE in river sediments were decreased by 46% and 17%, respectively, when nutshell biochar was applied at 500°C (Yang et al., 2018a). These findings indicate lower PAH bioavailability is correlated with pyrolysis at high temperatures, most likely due to the biochar’s increased aromaticity (Table 3).

Table 3

Adsorption of organic contaminates through biochar.

Organic contaminants

Feedstock and pyrolysis temperature (°C)

Sorption capacity/ removal %

Chloramphenicol Biphenol A

Bamboo eucalyptus wod, 380°C Potato peels, 400–800°C

233 mg/g 454.62 mg/g

Safranin T Salicylic acid, ibuprofen

Ramie biomass biochar, 500°C Pine wood, 425°C

Metribuzin SGBC Metribuzin MGBC Perfluorooctane sulfonate (PFOS) Carbamazepin e 2,4-Dichlorophenol

Switchgrass (SGBC), 450°C Magnetic switchgrass (MGBC), 450°C Cornstraw, 700°C

226.7 mg/g 22.70 mg/g 10.74 mg/g 223 m/g 205 m/g 169.30 m/g

Essandoh et al. (2017) Essandoh et al. (2017) Guo et al. (2017)

104.85–861.70 mg/L 99.95%

Chen et al. (2017) Kalderis et al. (2017) Kim et al. (2016)

Palm kernel shell, 660°C Macroalgae-derived biochar, 400–800°C Popular and conifer, 1200°C Wood chip, 1200°C Waste water sludge, 850°C

6.79 mg/L 9.22 mg/L 28.4 mg/L 30.2 mg/L 23.76–59.92 mg/g 5306.2 m/g 11.9 mg/kg 132 mg/kg 45 mg/L

Bamboo biochar, 300–700°C Sugarcane bagasse biochar, 400–800°C

61.68% 3000 mg/L

Digested bagasse biochar, 600°C

54.38 m/g 8.60 m/g

Corn straw, 300°C Corn straw/poplar leaf, 300–700°C

476.19 mg/g 155.9 m/g

Zhao et al. (2017a) Zhao et al. (2017a)

Corn straw, 300°C Saw dust, 600°C Saw dust, 600°C

476.19 mg/g 91.6% 10–25 mg/L

Zhao et al. (2017a, b) Zhou et al. (2017) Zhou et al. (2017)

Benzophene [BZP] Benzotriazole [BZT] Bisphenol A [BPA] 17 β-estradiol [E2] 4-Nitrotoluene Malachite green Phenanthrene Pentachlorophenol Sulfamethoxazole (SMX) N-nitrosodimethylamine Malachite green Sulfamethoxazole (SMX) Sulfapyridine (SPY) Bisphenol A (BPA) Dodecylbenzene sulfonic acid (DBSA) Biphenol A Tetraethyltin Tetracycline

Bamboo biochar, 500°C Sewage sludge/wood chip Hog fuel demolition waste, 350–600°C Pine chips, 300°C

References Ahmed et al. (2017) Arampatzidou and Deliyanni (2016) Cai et al. (2018) Essandoh et al. (2015)

Kim et al. (2016) Lin et al. (2018) Rao et al. (2017) Rao et al. (2017) Shimabuku et al. (2016) Chen et al. (2015) Vyavahare et al. (2018) Yao et al. (2018)

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Current Developments in Biotechnology and Bioengineering

4. Applications of biochar in polluted sediments The primary purpose of contaminated sediment treatment technologies is to eliminate or immobilize contaminants, thus decreasing pollutants’ danger to organisms or humans. Dredging and capping are two traditional sediment remediation techniques nowadays. Dredging treatments can help to extract contaminants from marine habitats more effectively. Dredging reduces labile metal fractions in soil, which improves the metal-binding ability of available sediment solids and delays metal leaching from solids (Chen et al., 2019a, b). On the other hand, it has some drawbacks, including the inability to remove all contaminants, the fact that new pollutants can move downward and contribute to pollution again, the loss of native benthic communities present in the soil, and the technique is costly (Akcil et al., 2015). Given the disadvantages of dredging, utilizing less expensive biochar for in situ capping to encapsulate contaminants is a feasible and growing option. The US Environmen€rstner and Apitz, tal Protection Agency has already successfully used capping technology (Fo 2007). By binding and adsorption, capping treatment effectively prevents contaminants from being exposed to sediments, creating a protective barrier (Zhang et al., 2017). Biochar and other materials such as zeolite, calcite and apatite can be combined with sediments to create a capping layer, or biochar can be combined with sediments to create a capping layer (Zhang et al., 2016). Three engineering techniques can be utilized to add biochar to polluted sediments; a biochar capping layer with gravel, zeolite, or quartz sand is utilized for low-flow water or closed reservoirs. Two permeable geotextile surfaces formed of the capping material are employed for high-flow water. For tiny water bodies, sediments mixed with biochar are employed as a capping layer. Alternative to this, the biochar was reported to exhibit adverse effects on aquatic organisms. The biotoxicity of pine needle derived biochar was studied by evaluating the endpoints such as cell growth, superoxide dismutase (SOD) content, reactive oxygen species, and chlorophyll-a (Chl-a) (Zhou et al., 2019). It was recorded that the free radicals in biochar led to the generation of intracellular reactive oxygen species in aquatic organisms. It was shown that the increase in the pyrolysis temperature from 200 to 500°C induced free radical formation. The role of biochar in the treatment of polluted sites is more detailed in the previous section.

4.1 Biochar coupling with capping materials Since sewage is the primary source of sediment contamination (Yang et al., 2014), retaining a barrier between sewage and sediment would help to reduce the sedimentation risk of pollution and prevent pollutants from releasing into the water from contaminated sediment (Ghosh et al., 2011). Biochar is an excellent sorbent for active capping because it can prevent pollutants from diffusing between water and sediment (Silvani et al., 2017). Traditional materials, including natural zeolite and sand, have been used to cap biochar, suitable for closed lakes and low-flow water. To minimize 4-chlorophenol and Cu2+concentrations, Zhang et al. (2017) introduced rice husk biochar into the sediments, supported by a quartz sand coating to avoid biochar movement. According to the findings, a thick layer of rice husk

Chapter 13 • Role of biochar in polyaromatic hydrocarbons remediation

377

biochar can prevent 4-chlorophenol and Cu2+ from releasing sediments via flocculation, adsorption, or microbial ingestion (Mishra et al., 2020b). Furthermore, it has been demonstrated that capping systems about the biochar layer have higher adsorption for NH4-N from water than the soil layer, preventing its spontaneous liberation from sediment (Zhu et al., 2019). The reactive core can be added to two permeable geotextile layers to make these clean materials more durable and avoid creating a biochar-sediment mixture. The use of a biodegradable geotextile layer allows the capping material to withstand high flow water and harsh weather (Wang et al., 2018).

4.2 Biochar mixed with polluted sediments Rather than capping with safe addition products, another method for immobilizing pollutants and reducing their bioavailability and accumulation in food chains is to combine biochar with sediment (Mishra et al., 2019). Long-term capping treatment technology is possible in the field when the capping treatment is combined with in situ bioremediation (Silvani et al., 2017). However, long-term capping treatment will minimize biochar’s remediation capacity (Lou et al., 2012; Raj et al., 2021). Capping material mixing is not practicable for large-scale sediment remediation, and mechanical mixing is only feasible for small-scale sediments or shallow water. Through combining biochar with contaminated sediments, pollutants’ toxicity, mobility, and bioavailability can be decreased (Liu et al., 2018). This technique is widely employed in laboratory-scale experiments for binding sediment contaminants (Wang et al., 2019).

5. Challenges in polyaromatic hydrocarbons remediation Over the years, various traditional and advanced oxidation methods have been used to remove PAHs from wastewater. Because of the complexities and harmful by-products, conventional approaches do not entirely remove PAHs. To assess and improve the efficiency, nonlinear methods such as Langmuir, Freundlich, Temkin, Redlich-Peterson, Webbers and Morris models, chemical, biological, electrochemical oxidation, and other existing processes can be studied. Nevertheless, the existing techniques have drawbacks and are inadequate in PAH degradation, considering the associated yield, removal efficiency, degradation of complex PAHs, ease of use, cost-effectiveness, analytical recovery, and performance evaluation. More research is required on PAHs elimination in wastewater using biochar. Combining biochar and other secondary systems with other treatment techniques is also needed to assess, analyze, and optimize operating conditions, study reaction intermediates and their performance, estimate degradation kinetics, and design efficient geometric configuration reactors and materials. Moreover, the application of the combined systems for PAH removal should be tactically designed to focus on the industrial scale (Varjani et al., 2017). Biochar is a renewable resource that can solve various environmental issues, including the removal of pollutants from soil, aqueous, and gaseous media. The use of activated

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Current Developments in Biotechnology and Bioengineering

biochar is another strategy to expand the application of biochar in pollutant removal. Nevertheless, there are still some challenges that need focus on to benefit from biochar-based remediation ultimately (Varjani et al., 2020). One of the significant problems is toxic compounds such as dioxetane, chlorinated hydrocarbons, PAHs, which may penetrate depending on the biomass used for production. In addition to the type of raw material, the evaluation and maintenance of proper production parameters such as pyrolysis temperature and pH also impact biochar characteristics, including sorption capability. Therefore, producing biochar for a specific process such as PAH remediation would require appropriate feedstock and process parameters. Moreover, before large-scale implementation of biochar-based pollutant removal, good economic advantages, environmental impacts, and life-cycle analysis are required. New techniques have been developed for biochar characterization and properties, but there is restrain on their proper utilization due to a lack of economic viability and accessibility estimation (Mishra et al., 2020a).

6. Conclusions and perspectives The development and implementation of strict guidelines for biochar production and quality assurance should be one primary focus in future studies. To better utilize the production biomass, methods and pyrolysis conditions, a database can be prepared, including biochar feedstock, production procedure, physicochemical properties, and associated function. There is still insufficient data regarding the efficient removal of PAH using biochar, and hence, it is likely inappropriate to conclude the overall process outcomes. Moreover, there is a need to investigate ways to modify biochar to improve its remediation efficiency by studying the surface chemistry of biochar, sorption and desorption studies, the exact mechanism behind the removal of contaminants, and life cycle assessment biochar. In the capping technique, biochar cost, efficiency, and capping layer design are the primary expenditure sources. Further improvement in capping technologies could optimally reduce the overall process costs and enhance PAH remediation. Also, an essential database for sediment remediation could be generated by long term monitoring of different capping layers at contaminated regions. To evaluate the practical feasibility of different biochar, toxicity investigations shall be focused to determine the adverse effects of biochar, if any, on the aquatic biota. Such studies and amendments would help in developing improved biochar-based systems and techniques for PAH removal from various sediments and consequently help in environment management.

References Adesra, A., Srivastava, V.K., Varjani, S., 2021. Valorization of dairy wastes: integrative approaches for value added products. Indian J. Microbiol. 61, 270–278. Ahmad, A., Singh, A.P., Khan, N., Chowdhary, P., Giri, B.S., Varjani, S., Chaturvedi, P., 2021. Bio-composite of Fe-sludge biochar immobilized with Bacillus sp. in packed column for bio-adsorption of methylene blue in a hybrid treatment system: isotherm and kinetic evaluation. Environ. Technol. Innov. 23, 101734. https://doi.org/10.1016/j.eti.2021.101734.

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Yao, Y., Zhang, Y., Gao, B., Chen, R., Wu, F., 2018. Removal of sulfamethoxazole (SMX) and sulfapyridine (SPY) from aqueous solutions by biochars derived from anaerobically digested bagasse. Environ. Sci. Pollut. Res. 25, 25659–25667. Zhang, C., Zhu, M., Zeng, G., Yu, Z., Cui, F., Yang, Z., Shen, L., 2016. Active capping technology: a new environmental remediation of contaminated sediment. Environ. Sci. Pollut. Res. 23, 4370–4386. Zhang, S., Tian, K., Jiang, S.-F., Jiang, H., 2017. Preventing the release of Cu2+ and 4-CP from contaminated sediments by employing a biochar capping treatment. Ind. Eng. Chem. Res. 56, 7730–7738. Zhao, N., Yang, X., Zhang, J., Zhu, L., Lv, Y., 2017a. Adsorption mechanisms of dodecylbenzene sulfonic acid by corn straw and poplar leaf biochars. Materials 10, 1119. Zhao, N., Zhao, C., Lv, Y., Zhang, W., Du, Y., Hao, Z., Zhang, J., 2017b. Adsorption and coadsorption mechanisms of Cr (VI) and organic contaminants on H3PO4 treated biochar. Chemosphere 186, 422–429. Zhou, Y., Liu, X., Xiang, Y., Wang, P., Zhang, J., Zhang, F., Wei, J., Luo, L., Lei, M., Tang, L., 2017. Modification of biochar derived from sawdust and its application in removal of tetracycline and copper from aqueous solution: adsorption mechanism and modelling. Bioresour. Technol. 245, 266–273. Zhou, W., Zhao, S., Tong, C., Chen, L., Yu, X., Yuan, T., Aimuzi, R., Luo, F., Tian, Y., Zhang, J., 2019. Dietary intake, drinking water ingestion and plasma perfluoroalkyl substances concentration in reproductive aged Chinese women. Environ. Int. https://doi.org/10.1016/j.envint.2019.03.075. Zhou, Y., Qin, S., Verma, S., Sarb, T., Sarsaiya, S., Ravindran, B., Liu, T., Sindhu, R., Patel, A.K., Binod, P., Varjani, S., Singhania, R.R., Zhang, Z., Awasthi, M.K., 2021. Production and beneficial impact of biochar for environmental application: a comprehensive review. Bioresour. Technol. 337, 125451. https://doi. org/10.1016/j.biortech.2021.125451. Zhu, Y., Tang, W., Jin, X., Shan, B., 2019. Using biochar capping to reduce nitrogen release from sediments in eutrophic lakes. Sci. Total Environ. 646, 93–104.
ska, A., Oleszczuk, P., Charmas, B., Skubiszewska-Zięba, J., Pasieczna-Patkowska, S., 2015. Effect of Zielin sewage sludge properties on the biochar characteristic. J. Anal. Appl. Pyrolysis 112, 201–213.

14 Environmental sustainability-based comparison for production, properties, and applications of biochar and hydrochar Misha Liu, Gajasinghe Arachchige Ganga Kavindi, and Zhongfang Lei FACULTY O F LI FE AND ENVI RONMENTAL SC IENC E S, UNIVERSI TY O F T SUKU BA, IB AR AK I, JA PA N

1. Introduction Biomass can be thermochemically transformed into carbon-rich solids including hydrochar and biochar materials by hydrothermal carbonization (HTC) and pyrolysis processes, respectively. Char is generally referred to the by-product from thermochemical conversion of organic solid wastes through pyrolysis, gasification, torrefaction, or HTC. The substantial yields of biochar, bio-oil, and syngas can be achieved through pyrolysis. In contrast, HTC process is commonly performed under thermal conditions with the presence of subcritical water, which can produce relatively higher yields of hydrochar and bio-oil with a low syngas yield (Ahmad et al., 2014). As the production process of hydrochar requires less activation energy to comply with the exothermic carbonization, HTC occurs at lower reaction temperatures than pyrolysis. At the same time, HTC is applicable for wet biomass without any pretreatment that contributes to the significant proportion of production cost by using the conventional pyrolysis process. Since water is served as the reaction medium instead of toxic and expensive solvents, the HTC process itself is a low-cost and environment-friendly alternative (Kambo and Dutta, 2015). On the other hand, the physicochemical properties of biochar are generally determined by several factors, in which the type of feedstock material plays a pivotal role in the chemical composition of biochar, and the reaction temperature controls the degradation process influencing the char properties. For instance, the microporosity and surface area of biochar may be increased at higher temperatures, while the low heating rate and prolonged holding time can maximize the biochar yield. In the case of hydrochar, however, the prolonged holding time of HTC may improve the specific surface area but reduce the hydrochar yield (Kavindi and Lei, 2019). In recent years, owing to the multiple physicochemical characteristics, biochar and hydrochar have gained much attention in the fields of environmental protection, energy Current Developments in Biotechnology and Bioengineering. https://doi.org/10.1016/B978-0-323-91873-2.00012-1 Copyright © 2023 Elsevier Inc. All rights reserved.

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and fuel, agronomy and analytical chemistry, which are mostly relating to low-cost adsorbents, soil amendment, carbon sequestration, bioenergy production, electrochemical devices, precursor of catalysts and anaerobic digestion (AD). The properties of biochar and hydrochar like large surface area, well-developed porous structure, abundant functional groups, and mineral components make them possible as candidate adsorbents for contaminants elimination. Further, the low density and porous structure of chars may make them applicable as a crop moisture conservation alternative (Bikbulatova et al., 2018). Based on an environmentally sustainable perspective, this chapter mainly aims to: (1) compare the production techniques for biochar and hydrochar and their corresponding conversion mechanisms; (2) discuss the physicochemical characteristics of the two types of chars; (3) review the predominant applications of biochar and hydrochar in diverse fields; and (4) identify the gaps and challenges in current research works and put forward the future perspectives.

2. Comparison between biochar and hydrochar production 2.1 Operational conditions Optimization of the operational conditions is the key to chars with desirable characteristics and properties. Moreover, the feasibility of the biomass conversion into char is totally dependent upon the production process and the quality of feedstock types. Thus, an in-depth discussion is provided in this section on the significant differences in operational conditions for the two biomass conversion processes.

2.1.1 Reaction temperature, pressure, and residence time Reaction temperature and residence time play a crucial role in the carbonization process, while both positively correlate with the carbonization and degradation of biomass. Biochar production is a thermochemical decomposition process, which can be classified as three major methods according to the changes in reaction temperature and heating rate in addition to residence time, namely, pyrolysis, gasification, and torrefaction (Table 1). Pyrolysis process for biochar production can be further divided into slow pyrolysis and fast pyrolysis. In the slow pyrolysis the heating rate is about 5–6°C/min coupled with prolonged residence time more than 1 h. As a result, the char yield can be increased due to the reduction of syngas percentage when compared to the fast pyrolysis (Kumar et al., 2020). However, the fast pyrolysis process can be realized at 200°C/min within a shorter reaction period, resulting in biochar with more excellent BET value and higher heating value (HHV) (Kambo and Dutta, 2015). Fast pyrolysis can be performed in different types of reactors such as circulating bed, rotating cone and fluidized bed (Zhang et al., 2019). Decomposition of wet biomass without any pretreatment can be realized through HTC processes that are divided into several subgroups based on the reaction temperature. At temperatures below 260°C, HTC or hydrothermal pretreatment process can yield a higher percentage of hydrochar; while under 250–350°C hydrothermal treatment or liquefaction

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Table 1

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Operation conditions for biochar and hydrochar production processes. Biochar Pyrolysis

Conversion Pretreatment Temperature (°C) Heating rate (°C/min) Residence time Pressure (MPa) Special conditions

Gasification

Thermochemical conversion Dry feedstock 300–700 600–900 5–200 50–100 0–6 h 10–20 s 1–2 1–2 No oxygen

Hydrochar Torrefaction

200–300 10–15 30 min–4 h 1–2

HTC

HTL

HTG

Hydrothermal carbonization No requirement of pretreatment 150–250 250–350 350–600 0–6 h 15–60 min 1–6 5–30 Subcritical water

< 60 min 25–100 Supercritical water

HTC, hydrothermal carbonization; HTL, hydrothermal liquification; HTG, hydrothermal gasification. Sources: Liu, Y.X., Yao, S.A., Wang, Y.Y., Lu, H.H., Brar, S.K., Yang, S.M., 2017. Bio- and hydrochars from rice straw and pig manure: intercomparison. Bioresour. Technol. 235, 332–337. Ahmad, M., Rajapaksha, A.U., Lim, J.E., Zhang, M., Bolan, N., Mohan, D., Vithanage, M., Lee, S.S., Ok, Y.S., 2014. Biochar as a sorbent for contaminant management in soil and water: a review. Chemosphere 99, 19–33. Kumar, A., Saini, K., Bhaskar, T., 2020. Hydochar and biochar: production, physicochemical properties and techno-economic analysis. Bioresour. Technol. 310, 123442. Kambo, H.S., Dutta, A., 2015. A comparative review of biochar and hydrochar in terms of production, physico-chemical properties and applications. Renew. Sust. Energ. Rev. 45, 359–378.

(HTL) process can achieve higher bio-oil yield. Moreover, the gaseous phase is increased beyond 350°C under the supercritical water condition that is termed as hydrothermal gasification (HTG) (Table 1). According to the summary in Table 1, the pyrolysis takes place under normal atmospheric pressure or slightly above (2 MPa) while HTC occurs at autogenerated pressure between 1 and 6 MPa. The pressure of the charring process can dramatically influence the reaction, which may also facilitate the rapid decomposition and carbonization. The reactor pressure usually tends to increase with the subcritical water pressure and the gas phase pressure, which is beneficial for carbonization process at low temperatures (Wiedner et al., 2013). However, during pyrolysis, secondary carbon formation and carbon percentage improvement are observed with energy densification under higher pressure (Kumar et al., 2020). Residence time is another important parameter to be considered in char production process when the yields of char, bio-oil and/or syngas are targeted. For instance, thermochemical conversion of biomass at higher temperature (600–1200°C) for a shorter residence time (10–20 s) is beneficial for production of more syngas with less char yield. On the other hand, in a slow pyrolysis process prolonging residence time at a low reaction temperature may enhance the char yield (Kambo and Dutta, 2015). On the contrary, HTC with prolonged residence time can reduce the char yield (Kavindi and Lei, 2019). Apart from the temperature, pressure and residence time, water plays a crucial role in the HTC process. Owing to the changes in physicochemical properties under higher temperature and pressure, water shows some similar behaviors as nonpolar solvents like acetone, methanol, and hexane under the operational conditions (Kritzer and Dinjus, 2001). Besides, the density of water reduces along with ionization into OH
and H3O+ during the process, which may facilitate the hydrolysis of biomass. However, water plays a negative

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role as it is energy intensive for predrying of wet biomass before biochar production process. Thus, HTC process is more beneficial for carbonization of wet feedstock with less activation energy while only dry biomass can be efficiently carbonized via pyrolysis (Kumar et al., 2020).

2.1.2 Feedstocks Biomass used as feedstock materials for char production mainly comprises 35%–50% cellulose, 15%–20% lignin, 20%–35% hemicellulose, and other substances including minerals, fats, and proteins (Nizamuddin et al., 2017). Thus, the physiochemical properties of the resultant char depend on the feedstock type. Lignin is the primary determinant of char yield, while cellulose and hemicellulose play a significant role in the enhancement of bio-oil content (Zhang et al., 2019). The degradation of lignin starts from 330°C (Raj et al., 2015), while during HTC process lignin degradation occurs at supercritical water state or above. However, both cellulose and hemicellulose decomposition takes place at 180°C and 160°C, respectively when HTC is applied, which starts above 200°C during pyrolysis (Kambo and Dutta, 2015; Zhang et al., 2019). Besides, macro/micromolecular compositions of the resultant char are dependent upon the composition of the feedstock used. For instance, the feedstock rich in minerals can yield a higher amount of char. Ahmad et al. (2014) reported that feedstocks like rice straw, wheat straw, coir dust or groundnut shell can produce comparatively higher char yields due to their higher K and Zn availability. Moreover, the feedstocks such as manure and urban wastes can be used to manufacture chars with higher nutrients content. In addition, determination of metals composition in biomass is important because they may influence biomass conversion and bioactivities of the produced char as well. For instance, metal cations such as K+, Mg2+, Ca2+, Cu2+, and Fe3+ available in rice straw, wheat straw and rice husk may exert different impacts on cellulase activities (Raj et al., 2015). On the other hand, the macromolecules in feedstocks generally undergo a hydrolysis process during HTC with furfural compounds in addition to the spherical carbonaceous nanoparticles being produced (Kambo and Dutta, 2015). Also, the conversion of nonapatite phosphate into apatite phosphate in the treated sewage sludge with resultant improved P bioavailability can be noticed, mainly attributable to the more formation of Al-P and Fe-P rather than Ca-P by HTC under neutral and acidic conditions (Yu et al., 2019).

2.2 Char modification Modification of biochar and hydrochar is a hot topic currently, aiming to achieve desirable chars with expected characteristics and properties. This can be realized either by feedstock modification with physical, chemical, and biological methods, or by process modification with changing operation temperature, residence time, air fractionation and/or combination of different processes. The properties or characteristics of biochar can be improved by surface modification, nanocomposite formation or structure modification including pore size and distribution.

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Surface oxidation is usually performed with chemicals like HNO3, H2O2 and/or KMnO4 to enhance dOH or dCOOH surface functional groups. Moreover, removal of impurities including some metal ions can help enhance char properties through acid modification (Wang and Wang, 2019). Nanocomposite formation possibly occurs either during the one-step production or coating with nanocomposite after the processing (Xiu et al., 2017). The biochar structure can be modified by chemical activation such as KOH, followed by some washing procedure resulting in enhanced surface area and pore structure; and the formation of precursor chemical during KOH activation can be removed by the washing procedure while leaving a porous structure that improves the surface area (Kambo and Dutta, 2015). Similarly, hydrochar modification can also be conducted by chemical, physical or biological means as a pretreatment or posttreatment of the char production process. Considering the nanocomposite incorporation, He et al. (2019) attempted the modification of tobacco stalk derived hydrochar with MgAl-LDH (layered double hydroxide) composite during the one-pot hydrochar production process, which possesses remarkably higher phosphate adsorption capacity when compared to the pristine samples.

2.3 Solvent extraction and char washing Minerals, bio-oils, or other impurities remaining in char can negatively affect the char performances by masking the reaction, which is same to biochar and hydrochar (Ahmad et al., 2014; Fang et al., 2018). Thus, char washing usually goes along with char modification to enhance the reactivity of biochar or hydrochar. Bio-oil remaining in hydrochar can be probably removed through the extraction process by solvents such as acetone and tetrahydrofuran. Removal of bio-oil attached to the interior surface of hydrochar has been identified by the interparticle diffuse model. Although C and H percentages of char were found to decrease, O and N percentages increased after tetrahydrofuran extraction; and surface area improvement was noticed after removal of hydrophobic compounds which masked the pores of char (Zhu et al., 2017). Some research emphasized the enhancement effect on oxygen-containing functional groups of biochar when methanol washing was followed (Jing et al., 2014). Similarly, the deashing process performed with acid-based washing procedures could enhance biochar’s surface area due to the removal of surface blockage. Also, bio-oil extraction after the HTL process can be carried out with solvent extraction and high energy gain, and the crude oil yield is highly dependent on the solvent selection (Watson et al., 2019).

2.4 Optimization of char production for sustainability concerns Energy demand is currently increasing at an exponential rate along with the population growth. Thus, the efficient and effective use of energy is a must. More importantly, the environmental pollution emissions at energy utilization should be addressed on top priority. Similarly, the efficiency and cleanness in the production of biochar and hydrochar should be critically evaluated. Hence, researchers paid more attention to the optimization

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and evaluation of conventional production processes while developing novel technologies to improve cost-effective production methods. Although direct biomass utilization is an attractive fuel production technology, it has several drawbacks including the requirement of pretreatment of feedstock, low density of energy, feedstock moisture, poor shelf life, inefficient energy utilization and release of pollutants such as tar (Ahmad et al., 2014; Kumar et al., 2020). Several methods can probably be applied for biochar production optimization, including the novel techniques such as microwave assisted pyrolysis, pyrolysis in presence of steam, wet pyrolysis, ammonia ambiance pyrolysis, and co-pyrolysis. Moreover, magnetic, acid or mineral modification, and electrochemical or carbon nanomaterial modification may enhance the efficiency and effectiveness of the engineered hydrochar under mild reaction conditions. Thus, these techniques more tend to be green production technologies when nontoxic reagents being applied to achieve safe handling of char beyond the operational boundaries (Wang et al., 2020b). Similarly, system optimization of hydrochar can be realized either by modifying the HTC method/approach with chemicals or by co-carbonization. For instance, according to Nizamuddin et al. (2018), homogenously heating process can take place in the microwave assisted carbonization where efficient energy may be supplied at the molecular level, resulting in a cost-effective process; furthermore, the microwave energy can be converted directly to internal energy as penetrating microwaves into the processed materials.

3. Comparison of structure and properties between biochar and hydrochar Char properties, greatly influenced by the production process and the composition of the feedstock materials, are the determinant factor of their potential applications such as sorbents, soil amendment and carbon sequestration. Although both biochar and hydrochar are possibly applied for the similar purposes, their reaction mechanisms are different. Hence, it is worth determining the characteristics of both biochar and hydrochar to identify their significant differences while selecting the ideal one for the most suitable application.

3.1 Char skeleton When the reaction temperature of pyrolysis process increases, a graphite-like structure can be observed along with the crystallization (Kambo and Dutta, 2015). Having a highly stable monolayer of carbon atoms, this graphene structure comprises electrical conductivity (EC) while resists to breakage either chemically or biologically. In addition, the aromatic carbon formation at higher temperature influences the resultant biochar’s adsorption capacity (Oliveira et al., 2017). Furthermore, the biochar porosity and crystallinity can be improved as a result of volatile matter removal at a higher reaction temperature (Qambrani et al., 2017). However, in HTC process sphere-shaped carbon particles

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instead of graphene sheets are formed, which are resulted from the particles arrangement as microspheres or tiny aggregates on the surface of hydrochar dominated by the alkyl moieties (Kambo and Dutta, 2015).

3.2 Physiochemical properties 3.2.1 Cleavage of bonds during char production Cellulose, hemicellulose, and lignin comprise alcohols, ketones, alkanes, esters, and other oxygen-containing functional groups. However, dOH and CdO bonds are common in cellulose while C]O bonds are abundant in hemicellulose. On the contrary, aromatic rings of C]C and CdOdC found in lignin are resistant to break down under low temperatures (Yang et al., 2007). During the thermochemical conversion under pyrolysis condition, hemicellulose degradation occurs at 200–260°C while cellulose degrades at 240–350°C, with lignin degradation starting from pyrolysis temperature >280°C (Qambrani et al., 2017). Hence, the cleavage of glycosidic bonds of xylan chain and formation of oligosaccharides can be observed during pyrolysis, which would then be further degraded into anhydro-Dxylopyranose species and other compounds (Kumar et al., 2020). Furthermore, aromatization, dehydration, and decarboxylation reactions are also involved in the production of syngas and bio-oil during pyrolysis (Zhang et al., 2019). Although the cleavage of lignin linkage and the formation of compounds binding with O and H containing functional groups have been observed, the actual mechanism behind the lignin decomposition in pyrolysis is still not clear (Zhang et al., 2019). When pyrolysis occurs at 400°C, electron donor or π electron-rich functional group formation can be detected, which may act as a receptor of electron-deficient groups (Wiedner et al., 2013). On the contrary, in HTC process cellulose chains can be firstly decomposed into oligomers, then forming glucose and fructose (Kumar et al., 2020). Similarly, hemicellulose can be converted to xylose and other furfurals which are observed as microspheres under a scanning electron microscope (SEM). However, the degradation of lignin is difficult, with some part being dissolved in water at 200°C. The dissolved lignin can be hydrolyzed into phenolic substances with phenolic char formed, while the remaining lignin that is not dissolved in water during HTC process can be converted into polyaromatic hydrochar (Zhang et al., 2019).

3.2.2 pH, electrical conductivity and cation exchange capacity The pH of biochar is comparatively alkaline with the increase in carbon percentage, while the cleavage of bonds associated with H and O occurs at a higher process temperature than HTC. Since the significant loss of H and O containing functional groups at reaction temperatures >500°C, biochar has higher aromaticity while reduced polarity (Oliveira et al., 2017). In contrast, the increase in H+ concentration under subcritical water condition (i.e., HTC) may facilitate the catalysis of organic compounds, so as to produce organic acids including lactic, acetic and formic acids as intermediate substances, which can reduce the hydrochar pH from HTC process (Wang et al., 2018).

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EC of char represents the total dissolved salts, which also provides information of the quality of available nutrients for electrochemical performance. The EC of both hydrochar and biochar increases with the increase of reaction temperature (Kavindi and Lei, 2019; Liu et al., 2017). Cation exchange capacity (CEC) can be used to measure the capacity of electrostatic interaction with positively charged molecules. Oxygen-containing functional groups, including hydroxyl and carboxyl groups on the char surface, mainly influence the CEC of char (Dieguez-Alonso et al., 2018). However, the data obtained by Takaya et al. (2016) indicate that the reaction temperature has limited impact on char CEC under the designed reaction temperatures for biochar (400–450°C and 600–650°C) and hydrochar (250°C), which is likely dependent on the feedstock type. Although solvent extraction of hydrochar can improve CEC by improving the surface porosity and avoiding pore blockage, biochar CEC is not significantly influenced by the extraction process (Takaya et al., 2016).

3.2.3 Bulk density The bulk density (BD) of hydrochar produced at higher process temperatures is lower than that of the feedstock, reflecting the improvement of char porosity (Kavindi and Lei, 2019). _ Chemerys and Baltrenait e_ (2017) reported that the biochar BD increased after being modified with MgCl2. In general, the char organic matter (OM) percentage decreases with the increase in temperature and holding time due to the more volatilization of organic compounds in the feedstock under a higher process temperature.

3.2.4 Moisture retention and hydrophobicity The physicochemical properties of char are closely associated with its moisture retention capacity, especially char porosity and surface functional groups. The polar functional groups on char surface have been reported to attract moisture through hydrogen bonds and electrostatic interaction (Dieguez-Alonso et al., 2018). There are three types of water available in biochar, i.e., freezable free water attached on larger pores and particle surface, freezable bound water in micropores, and nonfreezable bound water; the phase transition of freezable bound water takes place along with the distribution of pore size (Bikbulatova et al., 2018). As a result of restricted motion on the surface, water can remain as nonfreezing water on the surface while that available in submicropores may remain as freezable bound water due to the surface tension (Raj et al., 2018). The macropores of char can transport rather than retain moisture while micropores would mainly retain the moisture in the char (Liu et al., 2017). Besides, the secondary pores can be formed in biochar due to particle aggregation during the alternative wetting/drying process (Villagra-Mendoza and Horn, 2018). Hydrophobicity of biochar reduces at higher temperatures, resulting from the alternation of hydrophilic and hydrophobic surface functional groups during the process (Gray et al., 2014). On the other hand, the hydrophobicity of hydrochar may increase with the removal of hydroxyl groups during the charring process, which is important for the quality

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determination of solid fuels, easing the handling and storage of hydrochar due to the improved resistance to adsorption of hygroscopic moisture (Wang et al., 2018).

3.3 Elemental composition and proximate analysis According to the European Biochar Certificate (EBC, 2012) standards, carbon percentage of produced char is the crucial determinant of char quality with a threshold value of 50% C. However, the char with less than 50% C is called as pyrogenic carbonaceous matter. When the feedstock is rich in minerals, the resultant char has a high ash content with reduced C fraction (EBC, 2012). Volatile matter, ash and fixed carbon contents can be determined from proximate analysis, which are used to ensure the char quality for utilization (Dieguez-Alonso et al., 2018). The ash percentage of hydrochar is comparatively lower than that of biochar produced at slow pyrolysis because of the demineralization of biomass under HTC (Kambo and Dutta, 2015). However, the minerals composition of feedstocks is the key factor to ash percentage of the resultant char. O/C and H/C ratios in addition to volatile matter content can better indicate the carbonization degree of the charring process. According to the EBC (2012) standards, the biochar with H/C ratio