Formation, characteristics, and applications of environmentally persistent free radicals in biochars: A review

Bioresource Technology 281 (2019) 457–468

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Review

Formation, characteristics, and applications of environmentally persistent free radicals in biochars: A review

T

Xiuxiu Ruana,b, Yuqing Sunc, Weimeng Dua,b, Yuyuan Tanga,b, Qiang Liua,b, Zhanying Zhangd, ⁎ William Dohertyd, Ray L. Frostd, Guangren Qiana,b, Daniel C.W. Tsangc, a

School of Environmental and Chemical Engineering, Shanghai University, No.99 Shangda Road, Shanghai 200444, China Center of Green Urban Mining & Industry Ecology, Shanghai University, No.99 Shangda Road, Shanghai 200444, China c Department of Civil and Environmental Engineering, The Hong Kong Polytechnic University, Hung Hom, Kowloon, Hong Kong, China d Centre of Tropical Crops and Biocommodities, Queensland University of Technology, GPO Box 2434, Brisbane, Queensland 4001, Australia b

G R A P H I C A L A B S T R A C T

A R T I C LE I N FO

A B S T R A C T

Keywords: Stabilized radicals Engineered biochar Contaminant degradation Green/sustainable remediation Waste valorization/recycling

Due to abundant biomass and eco-friendliness, biochar is exemplified as one of the most promising candidates to mediate the degradation of environmental contaminants. Recently, environmentally persistent free radicals (EPFRs) have been detected in biochars, which can activate S2O82− or H2O2 to generate reactive oxygen species for effective degradation of organic and inorganic contaminants. Comprehending the formation mechanisms of EPFRs in biochars and their interactions with contaminants is indispensable to further develop their environmental applications, e.g., direct and indirect EPFR-mediated removal of organics/inorganics by biochars. With reference to the information of EPFRs in environmental matrices, this article critically reviews the formation mechanisms, characteristics, interactions, and environmental applications of EPFRs in biochars. Synthesis conditions and loading of metals/organics are considered as key parameters controlling their concentrations, types, and activities. This review provides new and important insights into the fate and emerging applications of surface-bound EPFRs in biochars.

1. Introduction Biochar, as a carbonaceous material derived from low-cost biomass residues, has captured extensive interests due to its promising applications in agricultural, environmental, and biorefinery activities

⁎

(Mohanty et al., 2018; O’Connor et al., 2018; Rajapaksha et al., 2016a; Xiong et al., 2017; You et al., 2017). In recent years, intensive attention has been drawn to the environmental use of biochar due to its promising potential for pH buffer (Rajapaksha et al., 2016b), nutrient/ water retention (Lee et al., 2018; Yang et al., 2019c), contaminant

Corresponding author. E-mail address: [email protected] (D.C.W. Tsang).

https://doi.org/10.1016/j.biortech.2019.02.105 Received 31 December 2018; Received in revised form 21 February 2019; Accepted 22 February 2019 Available online 23 February 2019 0960-8524/ © 2019 Elsevier Ltd. All rights reserved.

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as H2O2 or S2O82− (Qin et al., 2016). Semiquinone-type radicals in dissolved organic matter (DOM) from biochar were involved in As(III) oxidation (Dong et al., 2014). The biochar with EPFRs can also exhibit highly effective removal of Cr(VI) by reduction to Cr(III) for industrial wastewater treatment (Zhao et al., 2018; Zhong et al., 2018). Therefore, recent studies suggest emerging applications of biochar by utilizing its EPFRs as an effective catalyst for environmental remediation. These findings provide new insights into the significance of EPFRs on biochar. It is well known that PM plays an important role in the adverse human health effects, causing various diseases such as cardiopulmonary dysfunction, asthma, and cancer (Lin et al., 2007; Martin et al., 2011; Saravia et al., 2013; Yang et al., 2017a). EPFRs in PM rather generate surface-bound hydroxyl radicals than free hydroxyl radicals in aqueous solution in the absence of H2O2; the former can induce various types of cardiovascular and pulmonary disease (Khachatryan et al., 2014). The adverse impact of PM on human health is attributed to EPFRs-induced oxidative stress by ROS generation. For example, the detection of EPFRs and %OH in the biological fluids verified that PM generates ROS that can induce an oxidant injury and modulate toxic responses in biological tissues (Kelley et al., 2013). The generation of ROS by quinoid redox cycling can also be a formation mechanism of ROS from PM2.5 containing semiquinone-type radicals, similar to the toxicity of combustion products such as cigarette smoke (Squadrito et al., 2001). Researchers have found EPFRs in biochar inhibitory to plant germination and growth when biochar was used in soil remediation. Biochar addition as soil amendment has been reported to have positive effects on plant germination, growth, and yield (Alburquerque et al., 2013), whereas negative impact has also been documented when biochar induces ROS which can inhibit seed germination and retard growth of root and shoot (Ippolito et al., 2012; Fang et al., 2014; Liao et al., 2014). In light of the important findings contributed by over a decade of research on EPFRs in environmental matrices produced from combustion system, thermal treatment, and airborne PM (Vejerano et al., 2018), this article provides a critical review of EPFRs in biochars, ranging from sources, formation mechanisms, characteristics, to the associated environmental implications for fostering future research and emerging applications. The influences of synthesis conditions (i.e., temperature and time) and external metals on the characteristics of EPFRs in biochars are discussed in detail. The formation mechanisms of reactive species and the corresponding performance of contaminant removal are summarized. The current limitations and future opportunities for advanced investigations and new applications of biocharEPFRs are also suggested.

adsorbent (Supanchaiyamat et al., 2019; Yoon et al., 2017), energy production (Cho et al., 2018; Jung et al., 2017; You et al., 2018), and catalyst support (Cao et al., 2018a,b; 2019; Xiong et al., 2018; Yang et al., 2019d; Yu et al., 2019). Biochar can improve soil fertility via enhanced fertilizers retention and stimulated microbial community (Igalavithana et al., 2017; Yang et al., 2019e), increase soil carbon storage (El-Naggar et al., 2018a,b), reduce atmospheric greenhouse gas emissions (CO2, CH4, and N2O) (Lee et al., 2017; Thangarajan et al., 2018; Yoon et al., 2018), and immobilize organic/inorganic contaminants as a highly efficient and cost-effective sorbent (Luo et al., 2018; Yang et al., 2018a,b; 2019a,b). However, adsorption is a nondestructive process transferring pollutants from one phase to another (Fang et al., 2015b). The surface area, pore volume, and surface hydrophobicity of biochars are actually quite low compared with activated carbon (AC) or nanoparticles. Consequently, the observed high sorption capacity of organic/inorganic contaminants on biochars should not only depend on their physical properties. Environmentally persistent free radicals (EPFRs) with unpaired electrons can exist in ambient air for hours to months (Lomnicki et al., 2008; Vejerano et al., 2011; 2012a; 2012b; Gehling and Dellinger, 2013). In contrast to other free radicals, EPFRs are resonance-stabilized and bound to external or internal surfaces of solid particles, which can be measured by electron paramagnetic resonance spectroscopy (EPR) (Dellinger et al., 2007; Chen et al., 2018). Their lifetime under vacuum appears to be infinite, while they react with molecular oxygen resulting in decay with time in air (Gehling and Dellinger, 2013). EPFRs are categorized into three classes, i.e., oxygen-centered radicals, carboncentered radicals, and oxygenated carbon-centered radicals (Dela Cruz et al., 2012). Recognized formation mechanisms of these radicals are chemisorption and electron transfer. By virtue of EPR technique, Dellinger and colleagues have validated the common observation of EPFRs in combustion-generated particulate matters (PMs), sediments, and soils, which can stimulate the generation of reactive oxygen species (ROS) (Dela Cruz et al., 2011; 2012; Kiruri et al., 2013; Truong et al., 2010). It is well known that transition metals such as Fe2+ can induce ROS formation (Bondy et al., 1998; Murakami et al., 2006a–c; 2007; Rasmussen et al., 2007; See et al., 2007; Verma et al., 2010), and EPFRs have similar properties as transition metals. In general, 5,5-dimethyl-1-pyrroline-N-oxide (DMPO) is used to trap ROS, which is detected by EPR spectroscopy (Khachatryan and Dellinger, 2011; Khachatryan et al., 2011; 2014; Gehling et al., 2014). EPFRs of PM2.5 in aqueous suspension generated a significant amount of %OH without the addition of H2O2 (Khachatryan et al., 2011). The number of hydroxyl radicals induced by EPFRs (as indicated by DMPO-OH spin adduct concentration) was found to increase with longer reaction time, suggesting the ROS formation is a catalytic cycle during the process (Valavanidis et al., 2005; Khachatryan and Dellinger, 2011). During the hydroxyl radical formation, superoxide radicals form as an intermediate product (Khachatryan and Dellinger, 2011). Biochar, similar to PMs and airborne fine particles, is also a product derived from combustion. With an increasing pyrolysis temperature, the concentration of EPFRs on biochars increased (Liao et al., 2014). The generation of EPFRs can also be attributed to the external transition metals that chemically adsorb onto biomass and transfers electron from polymer to metal center during pyrolysis (Fang et al., 2014). Therefore, organic contaminants can be eliminated via other small molecular free radicals activated by EPFRs. For instance, EPFRs on biochars can activate hydrogen peroxide (H2O2) or oxygen (O2) to generate hydroxyl radicals (%OH), superoxide radical (O2%−), and/or activate persulfate (S2O82−) to produce sulfate radicals (SO4%−), which were shown to efficiently degrade organic contaminants such as chlorobiphenyl (Fang et al., 2014), phenolic compounds and polychlorinated biphenyls (Fang et al., 2015a), diethyl phthalate (Fang et al., 2015b), thiacloprid (Zhang et al., 2018), and bisphenol A (Ruan et al., 2018). Moreover, organic chemicals can also be directly degraded on biochar surface by macromolecular free radicals without addition of any radical-activators such

2. Sources of EPFRs in environmental matrices EPFRs are widespread in the environment due to the easy formation in the post-flame and cool-zone regions of combustion systems and other thermal conversion processes (Dela Cruz et al., 2012). In addition to PM, EPFRs are also found in contaminated soil (Dela Cruz et al., 2011; 2012), thermal treatments of plastic and hazardous waste (Valavanidis et al., 2008a,b), tar balls and pyrolysis of biodiesel at high temperatures (Kiruri et al., 2013; Mosonik et al., 2018).

2.1. PM PM floating in ambient air is mostly generated from flaring of hydrocarbons at refineries, vehicular exhaust emissions, and biomass burning (Fine et al., 2002; Kennedy, 2007). Airborne PM contains high concentrations of organics and metals such as copper and iron (Smith and Aust, 1997; Vicente et al., 2011), which are in favor of forming EPFRs. Organic molecules surface-bound to PM can readily form EPFRs in PM after thermal processes.

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2.2. Thermal treatment of plastic and hazardous waste

3.1. Generation of EPFRs during biochar synthesis

For combustion of plastic materials (polyvinyl chloride, polyethylene, polystyrene, polyethylene terephthalate, and polypropylene) at 600–750 °C, EPFRs are found in particulate smoke emissions, residue ash, and airborne soot (Valavanidis et al., 2008a,b), owing to co-existence of low concentrations of heavy metals, high concentrations of lithophilic elements, and polyaromatic hydrocarbons (PAHs). Researchers have also found carbon-centered EPFRs from different combustion systems (Tian et al., 2009).

During the pyrolysis process, cellulose, hemicellulose, and lignin as the main biomass constituents undergo different reaction pathways at various destructive pyrolysis temperatures of 300 °C, 300–400 °C, and 350–450 °C, respectively (Nzihou et al., 2013). Hemicellulose and cellulose are initially decomposed into oligosaccharides via depolymerization, followed by the cleavage of glucosidic bond to form different monomers and monomeric radicals (Zhang et al., 2013). In contrast, lignin with more complex structure undergoes a more intricate decomposition involving sequential reactions of free radicals (Kibet et al., 2012). Firstly, radicals are formed due to the homolytic cleavage of αand β-alkyl-aryl ether, CeC, and CeO bonds, which require lower energies to dissociate and generate the corresponding radicals (Kotake et al., 2014). Subsequently, these radicals can couple with other products or abstract hydrogen from other molecules. Finally, biochar is produced following a series of sequential reactions, i.e., dehydration, decarboxylation, aromatization, and intra-molecular condensation (Collard and Blin, 2014). Transition metals can further transfer electrons to phenolic lignin (Fig. 2), which is abundantly produced as phenol or quinone moieties during pyrolysis at high temperature, thus forming surface-bound EPFRs in biochar (Fang et al., 2014). To comprehend the EPFRs generation in biochars, Liao et al. (2014) conducted an in-situ observation of EPR signals during pyrolysis of corn stalk, rice, and wheat straws at 200 °C, identifying two noticeable stages with decreased intensities of EPR signals. The first one appeared before 30 min, where the breakdown of side chains might be involved in the EPFRs formation. These early produced outer-surface EPFRs would quickly react and dissipate, in accordance with the initially decreased intensity of the EPR signals. Continuous pyrolysis accumulated a large quantity of EPFRs on the limited surface area. The obvious decrease of EPR signals at 60–120 min might be due to the interactions between these EPFRs. A sharp increase of EPR signals was observed at 120–150 min during the cooling process, which might compress macromolecule structures from different directions, break up chemical bonds, and stimulate generation of additional EPFRs. Moreover, the sharp increase of EPFR concentrations was mostly attributed to oxygencentered EPFRs, which might have resulted from the cleavage of C-O bonds and/or oxygen incorporation into the broken CeC bonds. After 30-min cooling (i.e., 150 min after initiating the pyrolysis), the EPR intensities were relatively stabilized. These late-produced EPFRs originated from macromolecules rather than small molecules, and were difficult to move in the molecular chain to react with other EPFRs and/ or molecules due to the steric hindrance. Hydrochar is typically fabricated via hydrothermal carbonization (HTC) process performed at relatively low temperature ranges (i.e., 160–260 °C) and under autogenic pressure (i.e., 2–6 MPa) (Lang et al., 2019), which is dominated by ionic reactions with abundant hydronium ions (H3O+) and hydroxide ions (OH−) induced from water auto-ionization in subcritical conditions (Wang et al., 2018). The organic compounds with high molecular weight could be catalytically hydrolyzed to oligomers and then to monomers (Wang et al., 2016). The reactive monomeric radicals that are generated due to cleavage of some weak bonds in the biomass and hydrolysis products can undergo dehydration and fragmentation reactions to generate aromatic intermediates and other radicals (Chen et al., 2017). Hydrochar is then produced after polymerization, condensation, or aromatization (Liu et al., 2018a), while EPFRs are formed after stabilizing some of the unpaired electrons in the radicals by cyclic carbon π-system (DemirCakan et al., 2009).

2.3. Contaminated soils Soil contaminated by aromatic organic contaminants under lowtemperature thermal treatment can form EPFRs. Oxygen-rich condition would form more pentachlorophenoxyl radicals, which is oxygen-centered radicals and may persist for a long time (dela Cruz et al., 2012). Researchers found that soil and sediment contaminated with pentachlorophenol contained EPFRs (dela Cruz et al., 2011). The transformation of PAHs on transition-metal modified montmorillonite at room temperature resulted in EPFRs formation (Nwosu et al., 2016; Jia et al., 2018). Compared with other tested systems, EPFRs were more readily formed on Fe(III)-montmorillonite which had been contaminated by anthracene, and EPFRs including anthracene-based radical cations and oxygenated carbon-centered radicals were identified with the half-life of 38.5 days (Jia et al., 2016). The existence of EPFRs was also found in PAHs-contaminated soils of former coking sites, where the EPR signals of the soil samples were obtained with g = 2.0028–2.0036 (Jia et al., 2017). 2.4. Tar balls and pyrolysis of biodiesel EPFRs were detected in tar balls after the oil spill. The EPR spectra indicated two types of organic radicals, i.e., asphaltene radical species in crude oil (g = 2.0035) and radicals resulting from environmental transformations of crude oil (g = 2.0041–2.0047) (Kiruri et al., 2013). When croton megalocarpus biodiesel was pyrolyzed at 600 °C, EPFRs (g = 2.0024) were found with the half-life of 431 days (Mosonik et al., 2018). 3. Formation mechanisms of EPFRs in biochars The formation mechanisms of EPFRs and the affecting factors on the concentrations and types of EPFRs in biochars are still not fully elucidated owing to the highly variable synthesis methods, diverse precursors, and complex reaction processes. This review mainly focuses on the hydrothermal and pyrolysis processes to illustrate the major formation mechanisms of biochar-EPFRs based on the latest literature. Regarding the EPFRs formation in environment matrices, transition metals and substituted aromatics have been well recognized as key factors of EPFRs formation. Silica loaded with multiple chlorine or hydroxyl substituted aromatics and transition metals oxides (such as Fe2O3, CuO, ZnO, and NiO) can serve as a model particle for revealing the formation mechanisms EPFRs (Lomnicki et al., 2008; Vejerano et al., 2011; 2012a; 2012b; Kiruri et al., 2014; Yang et al., 2017b). In general, the most possible mechanisms include initial physisorption, chemisorption via elimination of water or hydrogen chloride, and then single electron transfer from substituted aromatics to the center of transition metals, which leads to the simultaneous reduction of metal and the formation of EPFRs (Dellinger et al., 2007). Transition metal accepts an electron and its valence changes from high to low during this process. Chlorine and hydroxyl substituted aromatics such as monochlorphenol tend to eliminate HCl and H2O at the same time (Vejerano et al., 2012b). As a result, the stability of EPFRs is attributed to the synergy of metals and aromatic compounds (Lomnicki et al., 2008).

3.2. Influence of synthesis conditions Fang et al. (2014 and 2015a) synthesized biochars from pine needles, wheat straw, and maize straw at different temperatures (300–700 °C) for various time (1–12 h). The results indicated that the 459

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Fig. 1. EPFRs generation in biochars synthesized via HTC process and the influence of external metals (). Adopted from Ruan et al., 2018

moieties via condensation and re-polymerization as temperature further increased (> 240 °C) (Funke and Ziegler, 2010), resulting in a pronounced reduction of the EPFRs content. The decrease in EPFRs concentration was possibly caused by condensation of aromatic rings and saturation of hydrogen due to aromatization at subcritical conditions (Landais et al., 1994). The EPFRs formation during HTC process was also time-dependent, as the peak area of the EPR signal decreased remarkably with an increase of residence time from 1 to 6 h. On one hand, a longer residence time might stimulate more intense degradation of the active moieties (e.g., PCs) in hydrochar responsible for EPFR formation. On the other hand, the formed EPFRs were decomposed and decayed in the HTC liquid with a prolonged period, probably due to the increasingly saturated biomass conversion with the steric hindrance at a higher condensation degree of aromatics (Demirbaş, 2000). Moreover, the EPR signal intensity decreased markedly with lower solid loading from 1:2.5 to 1:20. In other words, more EPFRs could be produced at a higher solid loading from the bond cleavage in the biomass polymers. The relatively smaller amount of water probably caused a weakened interaction between water and biomass components as well as an incomplete hydrolysis of straw. Longer residence time was required to reach equilibrium of biomass conversion at higher solid loading, resulting in a smaller proportion of fragment dissolved from biomass in HTC liquid. The saturation of EPFR-forming components due to hydrogen donation was suppressed, and the recombination of EPFR due to re-polymerization of dissolved fragments was also inhibited.

concentrations and types of the formed EPFRs were dependent on pyrolysis temperature and time. The optimized pyrolysis time for the generation of EPFRs at relatively low temperatures (300 and 400 °C) was 12 h, while it was 1 h at relatively high temperature (500, 600, and 700 °C) in the present study. The phenolic compounds (PCs) in biomass would be continuously decomposed with increasing pyrolysis temperature and time, resulting in reduced amount of PCs participating in the development of EPFRs. In addition, EPFRs could be further decayed as pyrolysis temperature and time continued to increase. In general, oxygen-centered EPFRs were the predominant species at relatively low pyrolysis temperature and short pyrolysis time, which were decayed or decomposed and transformed to carbon-centered EPFRs as temperature and time further increased. Yang et al. (2016) revealed that the EPFRs signals increased when pinewood was pyrolyzed up to 500 °C, while decreased significantly as the temperature increased to 700 °C. Qin et al. (2016) showed that pyrolysis of cow manure and rice husk at 700 °C generally eliminated all EPFRs due to the breakdown and reorganization of organic structures. Gao et al. (2018) studied the factors controlling the formation of EPFRs in hydrochar during hydrothermal conversion of rice straw. The EPR signal intensity increased with higher temperature from 180 °C to 240 °C and then remarkably decreased at 260 °C. At relatively low temperatures (< 240 °C), more EPFRs were generated due to newlyformed, active moieties in hydrochar during HTC process, which would probably participate in producing solid residues from the dissolved

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from rice husk via microwave induced pyrolysis at 1 kW for 30 min with EPFR concentration of 8.94 × 1017 spins g−1. Fang et al. (2014) synthesized biochars via pyrolysis of pine needle (P), wheat straw (W), and maize straw (M) at different temperatures (300–550 °C). The EPFR concentrations of biochars were 13.7 × 1018 spins g−1 for P550, 16.5 × 1018 spins g−1 for W400, 28.6 × 1018 spins g−1 for W500, 6.25 × 1018 spins g−1 for M400, and 30.2 × 1018 spins g−1 for M500. In another study of Fang’s group (Fang et al., 2015a), the EPFR concentration as a function of pyrolysis temperature and time were comprehensively investigated, which was increased rapidly for P300 and P400, while it decreased markedly for P500 and P600 under a prolonged pyrolysis time from 1 to 12 h. However, the EPR signal for P700 was below the detection limit. Qin et al. (2016 and 2017) developed biochars from cow manure (C) and rice husk (R) at 300–700 °C. Results showed that the relative order of EPFRs concentrations in the biochars was R-300 > C-300 > C-700 > R-700. Meanwhile, the EPFR concentration of R-300 to R 500 increased from 2.77 × 1018 spins g−1 to 17.10 × 1018 spins g−1, respectively; however, the EPFR content further decreased from 17.10 × 1018 spins g−1 to 0.16 × 1018 spins g−1 for R-500 to R700. Thus, biochar production at higher temperature (e.g., 500–700 °C) significantly reduced the EPR signals. The yield of EPFRs increases and then declines with increasing temperature because of the decomposition of radical structure (Dellinger et al., 2007; Vejerano et al., 2011). The g factor of corn straw biochar (300 °C) had value of 2.0040, which implied that the EPFRs belong to semiquinone-type oxygencentered radicals (Zhao et al., 2018). Based on Zhong’s results (Zhong et al., 2018), g factor of rice husk biochar was 2.0038, which was characteristic of carbon-centered EPFRs with an adjacent oxygen atom. Similarly, the g factors of biochars (300–550 °C) derived from pine needle, wheat straw, and maize straw ranged from 2.0028 to 2.0037, indicating that all of these EPFRs were carbon-centered radicals or carbon-centered radicals with an adjacent oxygen atom. The g-factor also varies with increasing temperature, indicating a change of types of EPFRs. Fang et al (2015a) concluded that oxygen-centered EPFRs converted to carbon-centered PFRs with an adjacent oxygen atom as the pyrolysis time and temperature increased. Qin et al (2016 and 2017) also determined that carbon-centered radicals with an adjacent oxygen atom were predominant in C-700 and R-700, while oxygen-centered radicals (e.g., semiquinone radicals) were dominant in C-300 and R300. Thus, the oxygen-centered radicals tend to decompose to carboncentered type at higher temperature. For example, semiquinone-type and phenoxyl-type radicals thermally decompose to form carbon-centered radicals above 700 °C (Dellinger et al., 2007). Semiquinone-type radicals can decompose to phenoxyl-type radicals (Vejerano et al., 2011). Through CO elimination, phenoxyl-type radicals can form cyclopentadienyl ketene compounds (Vončina and Šolmajer, 2002). In contrast, cyclopentadienyl radicals can survive at higher temperature due to greater resistance to thermal decomposition (Khachatryan et al., 2006).

3.3. Influence of external metals and PCs Fang et al. (2015a) suggested that both metal loading (i.e., Ni2+, Cu , Zn2+, and Fe3+) and PC (i.e., hydroquinone, catechol, and phenol) treatment on biomass (i.e., pine needle) favored the formation of EPFRs in biochar. The maximum concentrations of EPFRs were observed in the treatment of 0.1 mM metal or 5.0 mM PC, while their concentrations dropped with further increase in loadings. The conceivable reason was that transition metal ions presented a double-effect in the generation of EPFRs during biomass pyrolysis. The transition metal ions accept electrons from phenolic compounds and favor the formation of EPFRs at relatively low concentrations. However, the excess of transition metal ions would consume EPFRs on biochar due to the electron-shuttling effect of EPFRs accelerating the reduction of transition metal ions. Similarly, PCs exhibited a comparable ability to Fe3+ or Fe2+ consuming EPFRs via oxidation or reduction. Ruan et al. (2018) investigated the transformation of functional groups and EPFRs in the HTC of lignin (Fig. 1). During the HTC process, lignin primarily consisting of phenolic compounds was hydrolyzed into many intermediate products with hydroxy and carboxyl groups such as phenol, catechol, and hydroquinone, which were subsequently transformed into the hydrochars via polymerization and aromatization. The addition of FeCl3 promoted dehydration and decarboxylation and increased content of oxygen-containing functional groups, which could in turn form EPFRs with the chelated ferric ions. The electron transfer from the metal ions to the aromatic carbon structure contributed to the formation of phenoxyl- and semiquinone-type radicals. The phenoxy free radicals were further decomposed or reduced by carbon monoxide to form carbon-centered radicals. Meanwhile, the oxygen-centered and semiquinone radicals were formed from catechol and hydroquinone with several phenolic hydroxyls. 2+

4. Characteristics of EPFRs in biochars 4.1. Types of EPFRs EPFRs are a kind of stable and relatively unreactive radical, and these resonance-stabilized radicals are different from commonly shortlived radicals (hydroxyl radical, phenyl, vinyl, and methyl). The resonance-stabilized radicals can be detected by EPR, of which the gfactor values commonly used for identifying carbon-centered or oxygen-centered radicals (Valentin et al., 2006). The g-factor of EPFRs in environmental matrices change with metals and temperatures, which are summarized in Supplementary Material. Researchers typically categorize the EPFRs into three types, i.e., carbon-centered radicals (g < 2.003), carbon-centered radicals with an adjacent oxygen atom (g = 2.003–2.004), and oxygen-centered radicals (g > 2.004) (Dellinger et al., 2007). The oxygen-centered radicals are more stable in an atmospheric environment, whereas carboncentered radicals are more susceptible to oxidation in air. Semiquinone radicals (g > 2.0045) are oxygen-centered; phenoxyl radicals (g = 2.0030–2.0040) are oxygenated carbon-centered radicals; and cyclopentadienyls (g < 2.003) are carbon-centered radicals. Therefore, semiquinone-type radicals are found to be more resistant to react with molecular oxygen than cyclopentadienyls in ambient environment.

4.1.2. Influence of external metals Fang et al. (2015a) showed that the g-factors of EPFRs decreased from 2.0042 to 2.0032, 2.0039, 2.0036, and 2.0040 for 0.1 mM of Fe3+, Ni2+, Cu2+, and Zn2+ treatments, respectively, which suggested that metals loaded on the biomass changed the types of EPFRs in biochar. As suggested by Ruan et al. (2018), all of the 240 °C-Fe hydrochars were mainly composed of carbon-centered radicals with an adjacent oxygen atom and oxygen-centered radicals, which indicated that the Fe(III) addition promoted the formation of EPFRs such as semiquinone- and phenoxyl-type radicals. It is well known that the presence of transition metals is a significant factor to catalyze the formation of EPFRs and influence the types and yield of EPFRs. The abilities of the metal oxides to form EPFRs follow the order of oxidizing strength of the metal cations, i.e., Fe2O3 > ZnO > CuO > NiO (Yang et al., 2017b). Phenoxyl-type and

4.1.1. Influence of synthesis conditions According to recent studies, the EPFRs contents of fly ashes varied significantly from different sources. Feld-Cook et al. (2017) suggested that the presence of sulfates in some fly ashes might poison the active sites of transition metals and limit the formation of EPFRs. It has been reported that the type and yield of EPFRs in biochars rely on materials and temperatures as shown in Table 1. Zhao et al. (2018) calculated the EPFR concentration of 3.97 × 1018 spins g−1 in biochar derived from pyrolyzing corn straw at 300 °C. Zhong et al. (2018) prepared biochar 461

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Table 1 Characteristics of EPFRs formed in biochars. Material

Synthesis condition (°C, h)

Aromaticity (C/H)

Metal content (g kg−1)

SBET (m2 g−1)

EPFR Con. (10

Pine needles Wheat straw

Maize straw

Cow manure

P P P P P P P P

(350) (550) (300) (400) (500) (300) (400) (500)

P P P P P P P

(300, (700, (300, (400, (500, (600, (700,

spins g

−1

)

g factor

10.5 15.9 10.6 18.9 29.2 9.37 13.1 18.0

0.0342 0.0431 0.0366 0.0429 0.0464 0.0313 0.0373 0.0471

3.24 8.31 5.26 12.3 21.6 4.52 8.32 12.4

1.96 13.7 7.72 16.5 28.6 3.88 6.25 30.2

2.0034 2.0028 2.0036 2.0030 2.0029 2.0037 2.0031 2.0029

Fang et al., 2014

—

12.3 22.6 3.70

2.20 0.95 2.77 6.40 17.1 1.76 0.16

2.0046 2.0036 2.0039 2.0038 2.0034 2.0032 2.0032

Qin et al., 2016 and 2017

3.97

2.0040

Zhao et al., 2018

4.71

3.52 121 0.99 1.89 13.4 152 107

P (300, 2)

11.4

—

5.09

Rice husk

MW (1 kw, 0.5)

18.2

1.72

40.2

0.894

2.0038

Zhong et al., 2018

Pine needles (PN)

P P P P P P P P P P P P P P P P P P P P

9.61 12.8 14.7 17.5 — 16.7 18.8 21.6 23.7 — 21.6 25.1 34.6 39.0 — 23.9 28.2 36.0 43.0 —

Fe (1.5–2.2) Zn (0.061–0.075) Mn (0.11–0.19)

5.24 7.31 10.3 12.6 — 11.3 15.7 20.5 25.6 — 15.6 20.3 26.4 34.2 — 45.5 54.3 64.2 58.9 —

0.896 3.09 4.42 5.35 15.8 7.51 9.62 11.4 14.2 34.2 15.5 13.5 2.19 1.01 0.14 36.7 0.598 0.473 0.195 0.065

2.0048 2.0047 2.0045 2.0045 2.0041 2.0042 2.0042 2.0041 2.0039 2.0036 2.0038 2.0038 2.0037 2.0036 2.0032 2.0037 2.0035 2.0034 2.0034 2.0033

Fang et al., 2015a

PN + 0.1 mmol Ni PN + 2.0 mmol Ni PN + 0.1 mmol Cu PN + 2.0 mmol Cu PN + 0.1 mmol Zn PN + 2.0 mmol Zn PN + 0.1 mmol Fe PN + 2.0 mmol Fe PN + 1 mmol CT PN + 5 mmol CT PN + 20 mmol CT PN + 1 mmol PH PN + 5 mmol PH PN + 20 mmol PH PN + 1 mmol HQ PN + 5 mmol HQ PN + 20 mmol HQ

P (400, 2)

—

25.0–124

—

521 332 212 192 41.2 32.3 45.1 35.8 132 201 72.3 89.4 142 51.1 40.3 56.4 83.2

—

Lignin

H H H H H H H H

13.4 15.9 18.9 21.5 14.0 15.0 19.5 22.7

0.30 1.48 1.37 0.97 6.89 7.22 9.10 9.13

—

5.54 3.71 5.89 6.51 4.65 6.55 3.10 7.21

2.0037 2.0032 2.0037 2.0035 2.0035 2.0035 2.0041 2.0039

Rice husk

Corn straw

Lignin + Fe

(300, (300, (300, (300, (300, (400, (400, (400, (400, (400, (500, (500, (500, (500, (500, (600, (600, (600, (600, (600,

(240, (240, (240, (240, (240, (240, (240, (240,

4) 4) 4) 4) 4) 4) 4)

Reference 18

1) 2) 4) 8) 12) 1) 2) 4) 8) 12) 1) 2) 4) 8) 12) 1) 2) 4) 8) 12)

2) 4) 6) 8) 2) 4) 6) 8)

16.4 19.1 25.0 37.0 45.9

—

Ruan et al., 2018

Notes: P (pyrolysis), H (hydrolysis), MW (microwave), CT (catechol), PH (phenol), and HQ (hydroquinone).

radical (g = 2.0050–2.0065) (Vejerano et al., 2011). NiO produces phenoxyl-type radicals (g = 2.0029–2.0044) and semiquinone-type radicals (g = 2.0050–2.0081) (Vejerano et al., 2012b). The g-factor of EPFRs is slightly different among transition metals. The yield of EPFRs

semiquinone-type radicals are formed via catalysis by copper and zinc, while more than one type of organic radicals are formed on nickel and ferric iron. For instance, two EPFRs formed on Fe2O3 including a phenoxyl-type radical (g = 2.0024–2.0040) and a semiquinone-type 462

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4.3. Potential risks of EPFRs

formed on Fe2O3 is lower than CuO due to the higher oxidation potential of Fe2O3. The yield of EPFRs formed on NiO is the highest, which can be approximately 8 and 50 times higher than those on CuO and Fe2O3, respectively (Vejerano et al., 2012b). The concentration of transition metals would also affect the formation and persistence of EPFRs in particles (Kiruri et al., 2014). In the range of 0.5–15% CuO/SiO2 concentrations, the adsorption of 2,4-dichloro-1-naphthol on 5% CuO/SiO2 in PM produced the largest amount of EPFRs (Yang et al., 2017b). In other studies, comparing the range of 0.25–5% CuO/silica concentrations, the highest EPFRs yield was observed at 0.75–1% with semiquinone changing to chlorophenoxyl radicals. Changing the concentration may affect the metal clusters and its reactivity, thus the size of metal/metal oxide clusters plays a key role in determining the catalytic activity (Burda et al., 2005; Doyle et al., 2005; Bezemer et al., 2006; Wilson et al., 2006; Grass et al., 2008; Zhou et al., 2010). Compared to micrometer-sized metal oxides, nanoparticles could produce higher EPFRs concentrations (Yang et al., 2017b). Nevertheless, excessive metals may be reactive and consume EPFRs (Fang et al., 2015a).

EPFRs are widespread in the ambient environment and have links with various diseases such as influenza virus infection, acute/chronic pneumonia, and lung cancer owing to exposure to atmospheric PM (Lee et al., 2014; Chuang et al., 2017; Vejerano et al., 2018). Airborne PM, including vehicular exhaust and waste combustion, are apt to cause respiratory toxicity and dysfunction (Kreyling, 2004; Peters et al., 1997; Oberdürster, 2000). For instance, higher levels of PM2.5 increase the risk of myocardial infarction (Chang et al., 2013; Tsai and Yang, 2014) and exacerbation of respiratory disease such as asthma particularly for infant/child health because of incomplete development of immune system (Saravia et al., 2013). PM generated from wood smoke and biomass combustion contained organic molecules and metal particulates, which contribute to damaging health effects (Dellinger et al., 2001; de Kok et al., 2006; Valavanidis et al., 2008a,b; Danielsen et al., 2011). Semiquinone-type radicals are the most common EPFRs on PM (Pryor, 1998; Dellinger et al., 2000, 2001; Lingard et al., 2005; Valavanidis et al., 2005; 2008a,b), which can induce oxidative stress. The aerosol particles with EPFRs can result in the release of ROS inducing oxidative stress in the respiratory tract (Kennedy, 2007; Arangio et al., 2016). Airborne PM contains high concentrations of organics and metals such as copper and iron (Smith and Aust, 1997; Vicente et al., 2011), which are in favor of forming EPFRs. The toxicity of PM is determined by its size (Valavanidis et al., 2008a,b). Significant levels of hydroxyl radicals are generated by EPFRs in aqueous suspension, especially for those entrapped in the bulk of PM2.5 (Gehling et al., 2014). Hydroxyl radicals can damage cells and organs in the human body. The lifetime of EPFRs in PM2.5 range from 0.25 to 5028 days, thus EPFRs can be persistent in ambient air and radicals inhaled from PM2.5 can cause human health risk comparable to 0.4–0.9 cigarettes per day (Gehling and Dellinger, 2013). Liao et al. (2014) produced biochar from rice straw at 500 °C and used for germination and plant growth experiments with corn, wheat, and rice seeds. Strong %OH radicals were induced in solution by abundant biochar-EPFRs, which significantly inhibited germination, retarded root and shoot growth, and damaged plasma membrane. These adverse impacts were not obvious for biochars with low contents of EPFRs, while apparent hazards were observed after treatment of biochar with comparable concentrations of conventional contaminants (e.g., heavy metals and PAHs). Liu et al. (2018b) investigated the inhibitive effects of corncob biochars on urease-mediated urea hydrolysis, which was partially attributed to the released heavy metals and PAHs (∼20%) and primarily caused by the oxidative reactions with biocharEPFRs or their promoted generation of ROS (∼70%). Lieke et al. (2018) applied a model organism, Caenorhabditis elegans, to determine the neurotoxic effect of biochar (0–2000 mg C L−1) derived from pyrolyzing rice straw at 500 °C, which showed a hormetic effect on locomotion behavior. Biochar at high concentrations compromised defecation and recognition/response to chemical attractants. The observed neurotoxic effects could not be justified by any potential toxic chemicals with sufficiently high concentrations, but mainly due to the detected EPFRs in the biochar. Thus, the estimated neurotoxic effect of biochar was possibly attributed to the detrimental interaction of EPFRs with biotic macromolecules provoking oxidative stress responses accordingly. While biochars were perceived to exert potentially weak neurotoxicity to soil organisms, the relative risk of biochar-EPFRs should be addressed to ensure safe environmental applications.

4.2. Lifetime and decay of EPFRs The lifetimes of EPFRs are usually longer than free radicals, such that ERFRs stabilized on particle surfaces can be persistent in an atmospheric environment. The half-life of EPFRs is correlated with the standard reduction potential of the associated metal, and internal reaction with metal oxide is a pivotal factor affecting the stability of the radicals (Vejerano et al., 2012b). The metals to which the EPFRs are bound determine the stability and persistence of EPFRs. The half-lives of EPFRs on iron are longer than those on copper (Lomnicki et al., 2008; Vejerano et al., 2011). The half-lives of EPFRs on NiO ranging from 1.5 to 5.2 days are similar to those on Fe2O3 (Vejerano et al., 2012b). The half-lives of EPFRs on ZnO ranging from 3 to 73 days are the longest among the studied transition metals. In addition, the concentrations of transition metals also affect the half-lives of EPFRs, e.g., phenoxyl-type radicals on 0.5% CuO/silica and chlorophenoxyl-type radicals on 0.75% CuO/silica have the longest lifetime (Kiruri et al., 2014). The longest-lived EPFRs on ZnO can persist for nearly a year in ambient PM. This may be attributed to the enhanced interaction of EPFRs-metal complexes owing to small ionic radius and closed-shell structure of Zn2+ as well as the absence of crystal-field splitting energy. The stability of the radical is related to the reducibility of the metals and their ability to accept electron (Vejerano et al., 2012a). EPFRs may display fast decay, slow decay, or no decay on PM2.5, depending on their reaction with molecular oxygen in ambient air (Gehling and Dellinger, 2013). The fast decay and slow decay correspond to the decomposition of phenoxyl-type radicals and semiquinonetype radicals, respectively (Lomnicki et al., 2008; Vejerano et al., 2011; Vejerano et al., 2012a). No decay pattern is explained by the entrapment of radicals in the bulk of PM2.5 or restricted in a solid matrix (i.e., internal radicals), where the unpaired electron is delocalized over conjugated or aromatic bonds (Fraisse et al., 1993; Yordanov and Mladenova, 2007; Gehling and Dellinger, 2013). Liao et al. (2014) detected significantly increasing EPR signals with higher pyrolysis temperature in all the biochars synthesized from corn stalk, rice, and wheat straws, while the original biomass did not exhibit any detectable EPR signals. There was generally a less than 10% decrease of these signals after one month, which indicated the persistent nature of these detected stabilized EPFRs. However, the contribution by transition metals could be excluded in view of the low contents (∼mg kg−1) of all detected transition metals (e.g., Fe, Cu, Cr, Zn, and Pb). Thus, these biochar-EPFRs produced on the adjacent solid particle surface strongly interacted with the particles and were stabilized by steric hindrance.

5. Environmental applications of EPFRs in biochar Recent research utilizes the properties of EPFRs on biochar to catalyze different oxidants for generating ROS for organic contaminant degradation (Table 2). Fang et al. (2014) investigated the H2O2 activation by biochars (produced from pyrolyzing pine needles, wheat, and maize straw at 300 or 350 °C) for 2- chlorobiphenyl (2-CB) degradation. 463

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Table 2 Environmental applications of EPFRs in biochars. Biomass

Pyrolysis temperature (°C)

Dosage (g L−1)

Solution pH

Contaminant (Con.)

Oxidant

Removal efficiency (%) (Time, h)

Removal rate (kobs, min

(Molar ratio of contaminant to oxidant)

−1

Reference

)

Pine needles Wheat straw Maize straw

350 300 300

1

7.4 (10 mM PBS)

2-CB (10.6 μM)

H2O2 (1:943)

95 (2) 100 (2) 100 (2)

3.5 × 10−2 4.0 × 10−2 6.2 × 10−2

Fang et al., 2014

Pine needles

400

1

7.4 (10 mM PBS)

PCB-28 (3.9 μM)

Persulfate (1: 2051)

70 (4)

5.0 × 10−3

Fang et al., 2015a

Pine needles Wheat straw Maize straw

300

1

7.4 (10 mM PBS)

DEP (5 mg L−1)

O2 (1:28,594)

100 (24) 95 (24) 89 (24)

—

Fang et al., 2015b

Cow manure

300 700 300 700

333

—

1,3-D (15.3 mg L−1)

None

95.5 (384) 65 (384) 55 (384) 65 (384)

1.47 × 10−4 5.67 × 10−5 3.50 × 10−5 5.50 × 10−5

Qin et al., 2016

300 500 300 500

0.2

7.4 (10 mM PBS)

DEP (5 mg L−1)

UV (350–450 nm and centered at 365 nm, 2.3 × 10−5 Einstein cm−2 s−1)

58.9 72.3 52.3 60.9

4.50 × 10−1 6.48 × 10−1 3.78 × 10−1 4.74 × 10−1

Fang et al., 2017

Rice husk Pine needles Wheat straw

(2) (2) (2) (2)

Notes: PBS (phosphate buffered solution), 2-CB (chlorobiphenyl), PCB28 (polychlorinated biphenyl), DEP (diethyl phthalate), and 1,3-D (1,3-dichloropropene).

Fig. 2. Proposed mechanisms of EPFR formation and H2O2 activation by biochar synthesized via pyrolysis process (). Adopted from Fang et al., 2014

Results showed that 95–100% of 10.6 μM 2-CB was degraded by 10 mM H2O2 in the presence of 1.0 g L−1 biochars at pH 7.4 and 25 °C within 2 h. Therefore, H2O2 was effectively activated by biochar and •OH was predominantly produced via mechanism of single-electron transfer to degrade 2-CB (Fig. 2), which was supported by the linear correlations between EPFRs concentration, •OH generation, and 2-CB degradation rate (kobs). The degradation of 2-CB and decomposition of H2O2 were remarkably inhibited with prolonged reaction, probably due to slow degradation of sorbed 2-CB after the coverage or blocking of the reactive sites. Following the above study, Fang et al. (2015b) investigated %OH generation from biochar suspensions for diethyl phthalate (DEP) degradation in the presence of O2. Results showed that 89–100% of 5.0 mg L−1 DEP was degraded in biochar suspensions in the presence of O2 (0.2 mL min−1) with reaction time of 24 h. The generation of %OH could be induced by EPFRs in biochars at a consumption rate of ∼12 spins of EPFRs per molecule of %OH. The proposed mechanism was electron transfer from EPFRs in biochar to O2 producing the O2%− and % OH, which further reacted with EPFRs to produce %OH (Supplementary Material). Fang et al. (2015a) studied S2O82− (PS) activation by biochar for the degradation of polychlorinated biphenyl (PCB28). Results showed that 70% of PCB28 (3.9 μM) disappeared in biochar (1 g L−1)/ PS (8 mM) system after 4-h reaction. The activation of biochar by PS

was primarily controlled by the concentration and type of EPFRs, which was supported by linear correlations between EPFRs consumption and SO4%− formation, and between λ (λ = [SO4%− formation]/[EPFRs consumption]) and g-factors. Perceivably, PS activation by biochar was most likely due to the electron transfer from EPFRs to PS as follows (Fig. 3): (i) decomposition of S2O82− into SO4%− after electron transfer from biochar-EPFRs; (ii) production of O2%− after electron transfer from biochar-EPFRs to O2, and then reaction of O2%− with S2O82− producing SO4%−; and (iii) reaction of SO4%− with H2O or OH− to yield to %OH, and PCB28 degradation by SO4%− and %OH. In addition, organic contaminants may directly react with EPFRs in biochar without the addition of oxidants. Qin et al. (2016) developed biochars from cow manure and rice husk at 300 °C or 700 °C to investigate the catalytic degradation of 1,3-dichloropropene (1,3-D) in aqueous biochar slurry. However, the removal efficiencies (55–95.5%) and reaction rates (kobs = 0.35–1.47 × 10−4 min−1) were significantly lower than previous examples with oxidants (e.g., H2O2, O2, and S2O82−) even at a high biochar dosage (333 g L−1) and long reaction time (384 h). The 1,3-D degradation might be attributed to EPFRs and DOM in the biochar slurry. Meanwhile, photogeneration of ROS could be achieved by EPFRs in biochar under UV irradiation (Fang et al., 2017). Results showed that 52.3–72.3% DEP (20 mg L−1) degradation was observed in the presence 464

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Fig. 3. Proposed pathway for EPFR and SO4%− generation (). Adopted from Fang et al., 2015a

Fig. 4. Proposed framework for ROS formation from biochar suspension under light (Adopted from Fang et al., 2017).

of different biochars (0.2 g L−1) under UV light at pH 7.0 within 2-h reaction. It was concluded that biochar-derived DOM, quinone-like structure, and EPFRs played important roles in producing singlet oxygen (1O2) and %OH for DEP degradation. The mechanisms were proposed as follows (Fig. 4): (i) formation of 1O2 and %OH due to biochar-derived DOM via processes of light-induced energization and electron transfer; (ii) excitation of triplet-state quinone-like structure under light irradiation stimulating the 1O2 formation; (iii) UV-promoted EPFRs formation transferring electrons to O2 with the formation of O2%− and further yield of H2O2; and (iv) •OH production primarily via H2O2-involved pathways including EPFRs activation and photo-Fenton reaction. It is worth noting that EPFRs can be generated under UV irradiation (Li et al., 2014). In PAHs and ozone system, three radicals types are exhibited by the EPR spectra, two of which are semiquinone species and one is a PAHs-derived carbon-centered radical (Borrowman et al., 2016). Researchers degraded catechol under UV irradiation and found that phenol and semiquinone/quinine radicals were formed as a result of catechol degradation (Li et al., 2014), which may affect the residence time of organic contaminants in the environment. Apart from the catalytic degradation of organic contaminants, EPFRs on biochars were found to play an important role in inorganic removal from solutions in recent studies. Zhao et al. (2018) suggested that EPFRs on the surface of biochars were consumed during the removal of Cr(VI) in solution, implying that semiquinone-type EPFRs on the biochars directly donated electrons to reduce Cr(VI) into Cr(III) accompanied by formation of quinone groups (Supplementary

Material). Zhong et al. (2018) also illustrated that Cr(VI) could be reduced by carbon-centered EPFRs in the graphitic carbon-based compartment of biochar via electron transfer.

6. Current limitations and future outlook Even though the biochar production via HTC and pyrolysis processes has been extensively investigated, the formation mechanisms of biochar products rather than the surface-bound, resonance-stabilized EPFRs was primarily evaluated in most of previous studies. The influence of various precursors on the EPFRs generation is still not fully comprehended due to the complex and intertwined fundamental reactions during HTC and pyrolysis. There are deficient in-situ characterization techniques to trace and visualize the EPFRs formation processes during biochar synthesis. Both experimental and computational approaches should be improved and further efforts should be devoted to scrutinizing and tailoring the formation mechanisms of biochar-EPFRs, especially for biochars produced by different technologies and pyrolytic conditions. Current studies on biochar-EPFRs mostly focus on specific influencing factors on the concentrations and types of EPFRs in biochars. Nevertheless, the production of biochar derived from a wide range of biomass waste would require different pyrolysis temperatures and conditions. It is still challenging to design and tune the desirable properties of engineered biochar based on comprehensive understanding of the correlations between the biochar-EPFRs characteristics 465

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References

and the biochar synthesis conditions. The presence of multiple indigenous and exogenous transition metals and PCs can significantly alter the concentrations and types of EPFRs in biochar, but the presynthesis loading control is uneasy and the crosslinking catalytic effects on biomass pyrolysis are still not well controlled by design. Recent studies indicated that the catalytic activity of biochars can be further enhanced by surface modification via physical/chemical activation and hetero-atom doping. These engineering approaches could increase specific surface area and mesoporous structure, introduce active sites and functional groups, improve π-electron transfer, and alter electron density in the biochars. However, limited research studies have attempted to design the production of preferred concentrations and types of EPFRs for the sake of imparting superior biochar characteristics (e.g., surface functionality, catalytic reactivity, electron shuttling ability, hydrothermally stable carbon structure) for advanced applications in environmental, biorefinery, and energy industries. For instance, organic/inorganic contaminants can directly react with biochar-EPFRs under thermal activation or UV irradiation without the addition of other oxidants/reductants despite much lower reactivity in contrast with ROS. This presents a green and chemical-free approach in contrast to chemical-intensive advanced oxidation processes, but its potential has been largely overlooked in previous research. Meanwhile, the predominant mechanisms of electron transfer/shuttle from biochar to target contaminants are still controversial due to equivocal contribution of structural graphitization, surface functional groups, inherent/external metals, and biochar-EPFRs. Furthermore, many studies are committed to studying the ROS formation by lab-synthesized biochar under conditions that may be less relevant to the field-scale applications. Potential interactions of biochar-EPFRs with minerals, inorganic/organic compounds, and microorganisms may interfere with the degradation and fate of different contaminants in the environment. Thus, the actual selectivity and reactivity of biochar-EPFRs for environmental and other advanced applications have to be further validated.

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7. Conclusions Recent progress on the formation mechanisms and influencing factors of EPFRs with different types and concentrations in biochar were critically reviewed. Latest research utilizes the properties of EPFRs on biochar to catalyze H2O2 and S2O82− for generating ROS for contaminant degradation, although concerns remain about the biotoxicity of biochars due to surface-bound EPFRs and release of transformed organic/inorganic compounds. EPFRs on biochar may be further exploited to catalyze the generation of free radicals with or without chemical reagents under suitable operating systems and geochemical conditions, which potentially lead to wide-ranging environmental and other advanced applications.

Acknowledgement This project is supported by the National Natural Science Foundation of China (No. 21577085), Natural Science Fund Projects of Shanghai Municipal Science and Technology Commission (No. 14ZR1415600), Innovative Project of Shanghai Municipal Education Commission (No. 10YZ07), Key Subject of Shanghai Municipality (No. S30109), and Hong Kong Research Grants Council (E-PolyU503/17 and PolyU 15217818).

Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.biortech.2019.02.105. 466

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