Effects and mechanisms of biochar-microbe interactions in soil improvement and pollution remediation: A review

Environmental Pollution 227 (2017) 98e115

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Effects and mechanisms of biochar-microbe interactions in soil improvement and pollution remediation: A review* Xiaomin Zhu a, b, Baoliang Chen a, b, *, Lizhong Zhu a, b, Baoshan Xing c a

Department of Environmental Science, Zhejiang University, Hangzhou 310058, China Zhejiang Provincial Key Laboratory of Organic Pollution Process and Control, Hangzhou 310058, China c Stockbridge School of Agriculture, University of Massachusetts, Amherst, MA 01003, United States b

a r t i c l e i n f o

a b s t r a c t

Article history: Received 14 November 2016 Received in revised form 1 April 2017 Accepted 13 April 2017

Biochars have attracted tremendous attention due to their effects on soil improvement; they enhance carbon storage, soil fertility and quality, and contaminant (organic and heavy metal) immobilization and transformation. These effects could be achieved by modifying soil microbial habitats and (or) directly influencing microbial metabolisms, which together induce changes in microbial activity and microbial community structures. This review links microbial responses, including microbial activity, community structures and soil enzyme activities, with changes in soil properties caused by biochars. In particular, we summarized possible mechanisms that are involved in the effects that biochar-microbe interactions have on soil carbon sequestration and pollution remediation. Special attention has been paid to biochar effects on the formation and protection of soil aggregates, biochar adsorption of contaminants, biocharmediated transformation of soil contaminants by microorganisms, and biochar-facilitated electron transfer between microbial cells and contaminants and soil organic matter. Certain reactive organic compounds and heavy metals in biochar may induce toxicity to soil microorganisms. Adsorption and hydrolysis of signaling molecules by biochar interrupts microbial interspecific communications, potentially altering soil microbial community structures. Further research is urged to verify the proposed mechanisms involved in biochar-microbiota interactions for soil remediation and improvement. © 2017 Elsevier Ltd. All rights reserved.

Keywords: Biochar amendment Soil improvement Microbial community Carbon sequestration Contaminant mitigation Interaction mechanisms

1. Introduction For the purposes of soil remediation, biochar generally refers to a carbon-rich solid that is produced by the pyrolysis of biomass in oxygen-limited conditions (Beesley et al., 2011; Chen and Chen, 2009). Biochars are applied to soil due to their potential benefits for carbon sequestration, soil fertility, and contaminant immobilization (Cao et al., 2009; Chen et al., 2008a; Jeffery et al., 2015). The physiochemical properties of biochar are responsible for changes in soil character including changes in pH, nutrient maintenance, and water retention, which can induce heterogeneous responses in microbial species. This response can result in changes in microbial community structure and can consequently alter soil element

*

This paper has been recommended for acceptance by Dr. Hageman Kimberly Jill. * Corresponding author. Department of Environmental Science, Zhejiang University, Hangzhou 310058, China. E-mail addresses: [email protected] (X. Zhu), [email protected] (B. Chen), zlz@ zju.edu.cn (L. Zhu), [email protected] (B. Xing). http://dx.doi.org/10.1016/j.envpol.2017.04.032 0269-7491/© 2017 Elsevier Ltd. All rights reserved.

cycling and function (Biederman and Harpole, 2013; Lauber et al., 2009; Rousk et al., 2009, 2010). There are some components in biochar, including minerals, volatile organic compounds (VOCs), and free radicals (Spokas et al., 2011), that can potentially influence microbial activity, reshape the soil microbial community, and change the soil enzyme activity that catalyzes various key biogeochemical processes including soil organic matter turnover and elemental cycles (e.g., N, P, and S) (Paz-Ferreiro et al., 2014). Due to the various positive effects on the soil properties and microbes, biochars are considered effective agents for soil remediation. However, the variability among different types of biochar makes its effects on soil remediation quite unpredictable, and the specific mechanisms of biochar-microbe interactions are still unclear. The use of contaminant-degrading microbe inoculation (mycoremediation) together with biochar can enhance the biological degradation of pollutants (e.g., PAHs) (Chen and Ding, 2012; Chen et al., 2012a; Garcia-Delgado et al., 2015), providing a promising method for soil contaminant remediation. Such a process is considered a combination of the immobilization of the pollutants by the biochar and the further degradation of these pollutants by

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microbes. The adsorption of organic contaminants, toxic heavy metals, and hazardous anions (e.g., ClO
4 ) by biochar immobilizes them and prevents their leaching into the groundwater (Cao et al., 2009; Chen and Yuan, 2011; Chen et al., 2012d; Fang et al., 2014b; Jones et al., 2011; Yang et al., 2016b). Further transformation and detoxification of environmental pollutants by microbes as catalyzed by biochar have drawn increasing attention in recent studies (Dong et al., 2014; Oh et al., 2013). Persistent free radicals (PFRs) that are formed on biochar during thermal decomposition of the feedstocks activate reactive oxygen species (ROS) (Fang et al., 2015b; Kluepfel et al., 2014; Yu et al., 2015), and the electron transfer between biochar and microbial cells plays an important role in organic contaminant degradation and heavy metal transformation (Dong et al., 2014; Fang et al., 2015a; Yang et al., 2016a; Yu et al., 2015). Biochar can participate in soil processes such as organic matter decomposition as it takes part in the direct extracellular electron transfer (DEET) between soil organic matter (or soil minerals) and microbial cells, as well as in the direct interspecific electron transfer between microbial cells (DIET) (Chen et al., 2014; Fang et al., 2014a). The identification and quantification of the reactive components of biochar particles that are responsible for the electron transfer between the biochar and soil microbes are essential to investigate biochar-involved elemental cycling. The electron transfer between biochar particles and soil minerals, organic matter, pollutant molecules, and microbial cells, as well as in response of the microbial community to the reactive components of the biochar, is an emerging research field that seeks to further clarify the effects of biochar on soil biogeochemical processes. Several studies in the past decade extensively discussed the structure, physiochemical properties, and structure-function relationships of biochar with respect to the apparent effects of biochar on soil (Ahmad et al., 2014; Ameloot et al., 2013; Chen and Yuan, 2011; Chen et al., 2008a, 2012b; Lehmann et al., 2011; Warnock et al., 2007); however, an understanding of biocharmicrobe interactions is a research gap that would link biochar properties with many soil processes, e.g., carbon storage and contaminant degradation. With regard to the newly found electron transfer and free radical activation functions of biochar, the mechanisms of biochar effects can be more clearly identified (Fang et al., 2015b; Kappler et al., 2014). Therefore, this review aims to seek answers for the following questions: (1) What are the most essential physiochemical properties of biochar that influence the microbial activity and community, and how do these properties influence the microbes? (2) Mechanistically, how does the interaction of the soil microbes with different types of biochar affect soil carbon sequestration and contaminant dissipation? Answers to these questions are essential to identifying the cause of the heterogeneous effects of biochar in existing research and to develop the required understanding to accurately predict biochar effects, both of which are key challenges in this research field.

ergosterol extraction, quantitative real-time polymerase chain reaction (q-PCR), fluorescence in situ hybridization (FISH), phospholipid fatty acid quantitation (PLFA), molecular fingerprinting of 16S rRNA gene fragments including denaturing gradient gel electrophoresis (DGGE) and terminal restriction fragment length polymorphism (TRFLP), and high-throughput sequencing (also known as next-generation sequencing, or NGS) of soil microbial genes (Chen et al., 2013; Hale et al., 2014; Kolton et al., 2011; Mackie et al., 2015; Rousk et al., 2009). Changes in the relative abundances of Acidobacteria, Actinobacteria, Gemmatimonadetes, and Verrucomicrobia are frequently detected using high-throughput sequencing, under treatment with biochar (Mackie et al., 2015; Nielsen et al., 2014). With higher resolution to the species level, the metagenomics sequencing of microbial genes is able to realize function annotation reflected by the soil microbial community €ckel et al., 2004). Such a process is essential to structure changes (Ja explain the effects of biochar on soil remediation (Chen et al., 2013; Hale et al., 2014; Kolton et al., 2011; Mackie et al., 2015; Rousk et al., 2009). Since the mechanisms underlying biochar's effects on microbes and related soil functions and processes are still not quite clear, this review focuses on the synthesis of several possible mechanisms based on the published research. The influences of biochar on microbial activity are diverse and seven possible mechanisms are demonstrated in the central circle of Fig. 1 (from which points 1 to 3 can be classified into direct influences, and points 4 to 7 indirect influences): (1) biochar provides shelter for soil microbes with pore structures and surfaces (Quilliam et al., 2013a); (2) biochar supplies nutrients to soil microbes for their growth with those nutrients and ions adsorbed on biochar particles (Joseph et al., 2013); (3) biochar triggers potential toxicity with VOCs and environmentally persistent free radicals (Fang et al., 2014a); (4) biochar modifies microbial habitats by improving soil properties that are essential for microbial growth (including aeration conditions, water content, and pH) (Quilliam et al., 2013a); (5) biochar induces changes in enzyme activities that affect soil elemental cycles related to microbes (Lehmann et al., 2011; Yang et al., 2016b); (6) biochar interrupts microbial intra- and inter-specific communication between microbial cells via a combination of sorption and the hydrolysis of signaling molecules (Gao et al., 2016; Masiello et al., 2013); it should be noted that biochar may contain some molecules that can work as signals for microbial communication; and (7) biochar enhances the sorption and degradation of soil contaminants and reduces their bioavailability and toxicity to microbes (Beesley et al., 2010; Qin et al., 2013; Stefaniuk and Oleszczuk, 2016). The proposed mechanisms involved in biochar-microbe interactions need further experimental verification, and a research emphasis should be placed on the linkage between biochar-microbe interaction mechanisms and their environmental effects.

2. Driving mechanisms of biochar-microbe interaction in soil

One hypothesis of the benefits of biochar for microorganisms is that biochars can be shelters for microbes due to their pore structures. Biochars provide more habitable pore volume per unit volume than soil does (Quilliam et al., 2013a). Microbial living cells can attach on biochar surfaces; in such cases, biochars with large specific surface areas can provide habitats for microbes as well (Abit et al., 2012). However, the colonization of bacterial cells and fungal hyphae has spatial heterogeneity between the external and internal pores of biochar (Quilliam et al., 2013a). Different microbial colonization patterns on the surfaces and in the pores of biochar can be explained by three phenomena: 1) there is less nutrient accessibility in biochar pores than in natural soil pores, 2) the biochar pores can be blocked with soil organic matter (e.g., humic

Biochar affects the soil microbial activity and biomass, changes the soil bacteria to fungi ratio and soil enzyme activity, and reshapes the microbial community structure (Ahmad et al., 2016; George et al., 2012; Mackie et al., 2015; Nielsen et al., 2014; Rutigliano et al., 2014). Note that biochar application may significantly alter the microbial community structure even when it does not change the overall microbial activity and biomass. To clearly interpret the microbial responses to biochar application in soils, gene copy numbers can serve as a more sensitive parameter than microbial biomass (Chen et al., 2013). Various techniques are used to test microbial activity and community structure, including

2.1. Biochar provides shelter for microbes

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Fig. 1. Proposed mechanisms of biochar-microbe interactions and the environmental effects of biochar. The central circular area illustrates the interaction between biochar and microbes, while the enclosing four boxes represent the effects of their interaction on carbon sequestration, soil processes (elemental cycling), contaminant degradation, and plant growth. Interactions between the biochar and the microbes and its effects include: (1) biochar can act as a microbial shelter with its pore structure; (2) through sorption of nutrient cations via functional groups, biochar can improve soil CEC and maintain nutrients for microbial growth; (3) free radicals and VOCs on biochar can be toxic to some soil microbes, inhibit soilborne pathogens, and favor plant growth; (4) biochar can improve soil properties (e.g., pH, water content, and aeration conditions), and change the growth pattern of soil microbes; (5) biochar can adsorb enzyme molecules, influence soil enzyme activities and elemental cycles; (6) biochar can adsorb and enhance the hydrolysis of signaling molecules, and consequently interrupt microbial communication and alter microbial community structure; (7) biochar can enhance the sorption (via biochar surface functional groups) and degradation of soil contaminants (facilitated through electron transfer between biochar, microbes, and contaminants), which can reduce the toxicity of contaminants to soil microbes. The interactions between biochar and soil microbes can alter the microbial community and their metabolic pathways (which can be revealed by metagenomics analysis of microbial DNA sequencing), resulting in changed soil processes. There are interactions among different environmental effects as well.

Fig. 2. The interaction between microbial cells and biochar particles as observed with scanning electron microscopes (SEM). (A) Pine biochar inoculated with a co-culture of Geobacter metallireducens (rods) and Methanosarcina barkeri (spheres) for a short time period (20 days, scale bar ¼ 1 mm) (Chen et al., 2014) and (B) Field-aged biochars buried in agricultural soil for a long time period (3 years, scale bar ¼ 5 mm) (Quilliam et al., 2013a). (A) and (B) both show the attachment of microbial cells on biochar surfaces and pores. In (C), the mineral phases have entered the pores of wheat-straw biochar (Joseph et al., 2013). The SEM images show that the mineral phases have undergone considerable reactions that produce complex morphologies (Joseph et al., 2013).

acids), and 3) toxic substances, such as PAHs, may be present in biochar (especially in fresh biochar) (Kasozi et al., 2010; Quilliam et al., 2013a, 2013b). Microbial colonization on biochar surfaces and pores is also dependent on the biochar aging process, which can be considered a temporal heterogeneity (Quilliam et al., 2013a). Bacterial cells from co-cultures of Geobacter metallireducens and

Methanosarcina barkeri are able to attach themselves on biochar surfaces during a very short time period (20 days, as shown in Fig. 2A). The colonization of soil microbes (both fungal hyphae and single bacterial cells) on surfaces and in the pores of biochar can be improved by adjusting the aging periods of biochar (3 years, as shown in Fig. 2B) (Quilliam et al., 2013a).

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2.2. Biochar supplies nutrients to soil microbes Biochar contains a range of nutrients (e.g., K, Mg, Na, N, and P) (Chathurika et al., 2016; Rodriguez-Vila et al., 2016) and enriches soil nutrients via sorption that is due to its large surface area, high pore volume, and negative surface charge (Chen et al., 2012c). Cation exchange capacity (CEC) is a critical indicator of soil's ability to retain cationic nutrients and supply nutrients to support microbial activity. The improved soil CEC that results from biochar application reflects a higher nutrient retention capability and a lower nutrient loss through leaching, which is beneficial for soil microbial activity (Lehmann, 2007b), especially for microbes living in soils with low organic matter content (de Andrade et al., 2015; Laird et al., 2010; Mukherjee et al., 2011; Silber et al., 2010). The nutrient conditions of biochar (indicated by ash content) are largely determined by its feedstock and pyrolysis temperatures. Biochars produced from herbaceous material (crop residues) and manure generally have higher ash content than wood biochar and are consequently able to supply more nutrients (Fig. 3A) (Akhter et al., 2015). A clear rising curve between ash content with the pyrolysis temperatures has been found for crop residue biochar and manure biochar (Xu and Chen, 2013). The nutrients from biochar can be released with different rates into soils (Mukherjee and

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Zimmerman, 2013). In this case, biochar can act as a slow-release fertilizer, bringing long-term benefits to soil fertility and microbial growth. Notably, another reason that biochar supplies nutrients for soil microbes might be that it regulates soil microbial functions that are essential for nutrient cycling. For example, biochar can enhance the abundance of rhizobacteria that are able to transform organic S and P into bio-available forms, which further promotes the growth of Lolium perenne (Fox et al., 2014) and are reasonably suspected to promote the growth of other microbes that can only utilize inorganic S and P. Biochar stores and supplies nutrients to soil microbes via the sorption of nutrient cations and inorganic anions with its surface functional groups, especially oxygen-containing groups such as the carboxylate group (Fig. 4) (Chen et al., 2015; El-Naggar et al., 2015; Jien and Wang, 2013; Liang et al., 2006; Mukherjee et al., 2011; Yuan et al., 2016). Some studies showed that biochar maximized CEC at low and moderate pyrolysis temperatures, while to the contrary, some other studies showed that biochar CEC increased with the pyrolysis temperature (Lehmann, 2007a; Mukherjee et al., 2011; Yuan et al., 2011). The feedstock types and pyrolysis program parameters, including temperatures, heating rate, and holding time, primarily determine the biochar functional groups and consequently the capability of biochar to improve the soil CEC

Fig. 3. Biochar properties as function of pyrolysis temperature and feedstock types. (A) to (D) show the changes in ash content, specific surface area (SSA), pH, and volatile matter content (VM), respectively, with pyrolysis temperature. Different colors and shapes indicate different feedstock types, and linear (or nonlinear) regressions between biochar properties with pyrolysis temperatures are analyzed; the regression and its 95% confidence and 95% prediction bands are illustrated (feedstock types are not distinguished, analyzed with the software SigmaPlot). Positive relationships between pH and ash content with pyrolysis temperature are found, while negative relationships between VM content with pyrolysis temperature is found. An increased ash content with temperature is more obvious for manure biochar and crop residue biochar than for wood biochar (A), while an increased SSA with pyrolysis temperature is only found for wood biochar (B). “Other biochars” include distilled grain biochar and sugar cane bagasse biochar. Data are from Table S1.

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Fig. 4. Schematic diagram showing the roles of biochar functional groups (AFG ¼ acidic functional groups, SOM ¼ soil organic matter): (1) The AFG are responsible for the liming effect of biochar which modifies the soil microbial habitat; (2) the electrostatic attraction between the carboxyl groups of biochar with the nutrient cations effectively retains nutrients to ensure a nutrient supply to soil microbes and (3) to immobilize heavy metals, thus reducing heavy metal toxicity to microbial cells; (4) electrostatic attraction, as well as polar and non-polar organic attraction, of humic acid molecules can result in the adsorption of soil organic matter that is beneficial for carbon sequestration (further discussed in later chapter); (5) hydrogen bonding between eOH groups on biochar with oxygenated anions can adsorb inorganic anions to supply nutrients or reduce anion contaminant toxicity; (6) electron transfer to form free radicals on the biochar surface can facilitate organic contaminant degradation and heavy metal transformation (to be detailed discussed in later chapter) and can reduce contaminant toxicity to microbes.

(Table S1) (Chen et al., 2008a; Jeong et al., 2016; Mukherjee and Zimmerman, 2013; Yuan et al., 2011). The biochar CEC was pH dependent, increasing from low to neutral environmental pH values (Lehmann, 2007a), which indicated possible interactive effects between the changes in the pH and the CEC in soils with biochar. Additionally, interaction between biochars and soil minerals can be responsible for the long-term maintenance of the minerals during the aging of biochar (Novak et al., 2009). For example, a complex reaction between biochar (wheat straw) and soil clay as well as extra added chemical fertilizers (including CO(NH2)2, KCl and monoammonium phosphate) can result in some unusual morphologies (not identified) on biochar surfaces, as indicated in Fig. 2C (Joseph et al., 2013). From the above results, the various aspects of biochar preparation including feedstock types, pyrolysis temperatures, the aging period, and the relationships between various biochar properties (e.g., pH and CEC) should be fully considered during the field application of biochar to improve the nutrient supply to soil microbes. The aromaticity of biochar is responsible for its recalcitrant nature against microbial decomposition (reflected by the atomic H/ C and O/C ratios), and some fractions of biochar can also function as a carbon source for soil microbes (the quality of biochar as a carbon source is indicated by its C/N ratio, Table S1, Fig. S1) (Demisie et al., 2014). Biochars are typically low in available carbon for microbial utilization because they contain higher C/N ratios than their feedstocks and are thus difficult to degrade by microorganisms because of the lack in N supply. As a consequence, biochar amendment

increases the recalcitrant SOC pool, enhancing soil carbon sequestration (Yanardag et al., 2015). Bacteria and fungi have their own preferences for different carbon sources and have different tolerances to changes in environmental factors, such as pH and water condition (Rousk et al., 2009; Zhang et al., 2015). Compared with bacteria, fungi are able to colonize soil macroaggregations (>200 mm) that contain a higher SOC, ratio of C/N, and total N (Zhang et al., 2015). This ability may be due to the advantages of the living strategy of fungi; specifically, the translocation of nutrients and water within a continuous hyphal network enables fungi to colonize poor carbon sources with a high C/N, such as biochar (Ascough et al., 2010). Therefore, biochar application that promotes the formation of macroaggregates should favor fungi growth rather than bacteria under the same environmental condition. Both atomic H/C and O/C ratios commonly decrease with increased pyrolysis temperature, indicating a more intensive aromatic and nonpolar structure of the biochar (Fig. S1) (Hale et al., 2015; Xiao et al., 2016). Biochar C/N ratios are mainly dependent on feedstock types; for example, the C/N of wood biochar is normally higher than that of manure biochar (Cantrell et al., 2012; Chen et al., 2008a). In addition, the increased pyrolysis temperature can cause elevated C/N in some feedstocks due to the condensed aromatic carbon structure (Table S1, Fig. S1B) (Cantrell et al., 2012; Spokas et al., 2011). Therefore, the pyrolysis temperature and feedstock type are useful when considering the stability of biochar C against microbial decomposition in practical application.

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2.3. Potential toxicity of biochar to microbial cells Certain compounds in biochar are known as microbial inhibitors, including benzene (the dominant product of pyrolysis during glowing combustion of charcoal), methoxyphenols and phenols (the products of pyrolysis of hemicelluloses and lignin), carboxylic acids, ketones, furans (that are generally presented as the sorbed VOCs on biochar), and PAHs (Ghidotti et al., 2017; Lyu et al., 2016; Spokas et al., 2011). Such compounds can be identified from water or organic solvent extractions of biochar. Organic solvent extracts from biochar contain organic compounds from many classes, including n-alkanoic acids, hydroxyl and acetoxy acids, benzoic acids, diols, triols, and phenols, while water extracts from biochar contain dicarboxylic acids, aromatic organic acids and polyol along with hydroxy acids, n-alkanoic acids, and benzoic acids (Table 1) (Graber et al., 2010). Biochar VOCs vary with feedstock type, elemental composition, pyrolysis temperature, and heating procedures and conditions (Spokas et al., 2011). Higher concentrations and toxicities of PAHs and polychlorinated dioxins and furans (PCDD/DF) are found in biochar generated at moderate temperatures (300 and 400  C) than at high temperatures (>400  C) (Lyu et al., 2016). The diversity of composition for the VOCs species sorbed on biochar can be the main contributing factor to the various responses of soil microbial activity to biochar. Although volatile organic compounds (VOCs) in fresh biochar can support the survival of some microbes (such as Bacillus mucilaginosus) as a carbon source (Sun et al., 2015), they can induce potential toxicity to microbes (not species-specific) when present in high concentrations (especially for some low molecular weight oxygenated VOCs, including acids, alcohols, and carbonyls) (Ennis et al., 2012; Spokas et al., 2011). As microbial inhibitors, VOCs can also induce direct toxicity to microbial soil pathogens, thus benefitting plant growth (Graber et al., 2010). However, the inhibition may not be very specific to pathogens, and the alteration of the microbial community needs further investigation to evaluate the environmental benefits and risks of biochar application. Persistent free radicals (PFRs) that are generated and stabilized during the pyrolysis of biochar, including semiquinones, phenoxyls, cyclopentadienyls, and phenols, can also induce toxicity to microbes (Truong et al., 2010). The formation of free radicals on biochar includes replacing surface functional groups on metal oxides with aromatics while metal oxides are being reduced (Fig. 5I) (Balakrishna et al., 2009; Doong et al., 2014; Fang et al., 2014a) and breaking chemical bonds in macromolecules (Fig. 5II) (Dellinger et al., 2007; Liao et al., 2014). The PFRs are abundant (approximately 1017 spins g
1 pine needle biochar) and depends on the pyrolysis temperature and time (Fang et al., 2015a). The adverse impacts initiated by biochar PFRs were previously misidentified because of the lack of proper methods to separate those effects caused by free radicals from those caused by co-existing molecular pollutants; they have yet to be taken into account in current research (Liao et al., 2014; Truong et al., 2010). Free radicals can induce oxidative stress in living microbial cells, reduce cellular glutathione (GSH), glutathione peroxidase (GPx), and superoxide dismutase (SOD) levels, and decrease cell membrane integrity with the generation of reactive oxygen species (ROS) such as hydroxyl radicals (OH), the superoxide radical anion (O
2 ), and hydrogen peroxide (H2O2) (Balakrishna et al., 2009; Dellinger et al., 2007; Liao et al., 2014). Free radicals interfere with cytochrome P450s (e.g., CYP1A2, a member of the ubiquitous superfamily of enzymes that catalyze the mixed-function oxygenation of both endogenous and foreign compounds) and competitively inhibit the metabolism of exogenous organic substrates (Reed et al., 2015). Semiquinone radicals (QH) are one type of free radicals that are found in combustion-generated ultrafine

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particulate matter; these radicals can reduce oxygen to form superoxide, which then forms H2O2 that can initiate Fenton reactions in the presence of transition metal ions such as Fe2þ and Cu2þ (which are ubiquitous in biological systems). These Fenton reactions generate OH that causes DNA strand breaks, which lead to DNA damage (Dellinger et al., 2001). On the other hand, free radicals from biochar can play an important role in the degradation of organic pollutants with the ROS induced by the biochar, especially the strongest species: the hydroxyl radical (OH) and the sulfate radical anion (SO
4 ) (Fang et al., 2015a, 2015b), which will be discussed later. Both the potential toxicity of biochar free radicals on soil microorganisms and the positive effects of biochar free radicals on pollutant degradation are two sides of the coin to be considered in the evaluation of the environmental effects of biochar. 2.4. Biochar modifies microbial habitats Biochar can modify microbial habitats by improving the soil's physical properties. Biochar porosity can decrease soil bulk density, improve soil aeration condition (Abel et al., 2013), and control the transport of soil microbes in biochar amended soil (Abit et al., 2012). Biochar can increase the available water content that influences nutrient accessibility to microbial cells (Abel et al., 2013). In addition, the biochar can increase water content at the permanent wilting point, which indicates the capability of biochar, via its high porosity, to store water unavailable to plants; keeping water in this manner is beneficial, especially in sandy and degraded soils (Abel et al., 2013). Moreover, the improved water retention capacity means that there is a greater capability of the soil to hold water against dry-wet cycles in the natural environment, which can favor the maintenance of a stable microbial activity (Liang et al., 2014). The pyrolysis parameters (mainly temperature, heating rate and time) and feedstock compositions (e.g., lignin and lipid concentrations) of the preparation of biochar control its porosity, carbon stability, and the surface adsorption of nutrients (Abit et al., 2012; Cantrell et al., 2012; Chen et al., 2008a). Although an increasing trend of specific surface area with pyrolysis temperatures is found for wood biochar, this relationship is not found for biochars generated from other feedstock types such as manure and crop residue, indicating the extraordinarily important role of feedstock type on the formation of large SSAs (Table S1, Fig. 3B) (Chen et al., 2008a; Wang et al., 2015). The role of biochar in improving soil properties and modifying microbial habitats can be dependent on the feedstock types and pyrolysis procedures used in making biochar. Biochar can be an effective liming agent to neutralize soil pH (Yuan et al., 2011). An increase of the soil pH by 0.2e0.3 units after biochar application can be the main factor affecting the soil microbial community, in contrast to other chemical variables in soil such as the total C content, total N content, electrical conductivity þ (EC), and NO
3 and NH4 concentrations (Nielsen et al., 2014). An increase in the soil pH and a decrease in the toxicity of exchangeable Al in acidic soils with biochar amendment (Qian et al., 2013) increase the bacteria abundance in the pH range from 4 to 7; the abundances of bacteria are positively correlated with soil pH (Leal et al., 2015; Rousk et al., 2010). Fungi and bacteria have different sensitivities to soil pH (Rousk et al., 2009), and bacteria are generally more tolerant to a narrower range of pH and are more sensitive to soil pH changes than fungi are (Rousk et al., 2010). As a result, there are likely to be different responses from bacteria and fungi to biochar-induced changes in soil pH, and the microbial community of fungi and bacteria may react in different ways to biochar-induced changes in soil pH, which can result in an alteration of the overall microbial community structure. A pyrosequencing analysis of the soil bacterial community showed

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Table 1 Organic compounds frequently detected from various biochar types as VOCs, organic solvent extracts, or water extracts of biochar.

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significant correlation between the soil pH and the soil bacterial community composition. The relative abundances of Acidobacteria, Actinobacteria, and Bacteroidetes were largely driven by the range of soil pH values (3.5e9.0) at the continental scale from North to South America (Lauber et al., 2009). Increasing the soil pH with biochar application altered the abundance, diversity, and composition of nitrifying bacteria in soil and changed soil nitrification as a

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consequence (Zhang et al., 2017). Pyrolysis temperature dominantly determined biochar pH values, which usually rise with increased pyrolysis temperature for various feedstocks (Fig. 3C). This trend is mainly attributed to the decreased number of acidic functional groups (AFG) and decreased volatile matter (VM) content (Fig. 3D) (Mukherjee et al., 2011), yet there is an increased alkalinity (due to carbonates and alkaline ash content) of biochar

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Fig. 5. Schematic of the mechanisms of PFR formation and free radical generation on biochar, including: I) the interaction between organic compounds containing oxygenous functional groups and metal oxidation, and II) the breaking of chemical bonds in macromolecules during heating and cooling. The figure is modified from the reference (Fang et al., 2014a; Liao et al., 2014).

with increased pyrolysis temperature (Smebye et al., 2016; Yuan et al., 2011). Therefore, high temperature biochars are normally more effective at improving the soil pH (Cantrell et al., 2012; Cao and Harris, 2010; Mukherjee et al., 2014; Yuan et al., 2011).

2.5. Biochar alters soil enzyme activity Most of the elemental turnover in soil, which determines the nutrient bioavailability and includes turnover of C, N, P, and S, is catalyzed by enzymes. Soil enzymatic activities respond faster than other soil variables to soil management and disturbance introduced into soil and are therefore indicators of biological changes and soil quality (Bandick and Dick, 1999). Enzyme activities respond to biochar application in various ways, mainly depending on the types of enzymes and biochar, the biochar application rate, and soil properties (Table S2). The decreased microbial abundance and activity of soil enzymes in relation to organic matter decomposition can enhance C sequestration (Luo and Gu, 2016). Possible mechanisms involved in the influence of biochar on enzyme activity include but may not be limited to: (1) biochar adsorbs extracellular enzyme molecules and/or substrates on the surface or blocks the reaction site of enzymes (Bailey et al., 2011) in a way regulates their apparent affinity with substrates (Paz-Ferreiro et al., 2015); (2) biochar influences enzyme activity with changes in soil physiochemical properties (especially pH) (Zimmerman and Ahn, 2010); and (3) biochar releases some small molecules that are speculated to act as allosteric regulators or inhibitors of specific enzymes (such as a possible up-regulation of beNeacetylglucosaminidase activity with ethylene) (Bailey et al., 2011).

The sorption (binding) of enzymes on biochar and soil organic matter can change the kinetic properties of enzyme activity (Lammirato et al., 2011; Nannipieri et al., 2012) and therefore is the most important mechanism regulating the soil enzyme activity (Zimmerman and Ahn, 2010). The adsorption efficiency of the enzyme and substrates depends on the biochar structure: adsorption of enzyme molecules on biochar surfaces is considered to be driven by non-coulombic forces between the uncharged regions of the protein and the uncharged regions of the biochar surface, and the adsorption of small molecular polar substrates (e.g., disaccharide) on charred fractions (especially activated carbon) is stabilized via hydrogen bonding to polar surface groups (e.g., COOH, SO4H, PO4H) on the sorbents (Lammirato et al., 2011). Changes in surface functional groups in aged biochar alter the adsorption capacity of enzyme and substrates, thus affecting enzyme activity (Gibson et al., 2016). For example, the oxidation of aromatic carbon and the introduction of aliphatic C-H groups on biochar during the abiotic aging process improved laccase and peroxidase activities and enhanced fungal respiration (Gibson et al., 2016). Biochar can reduce the activation energy (Ea, related to the temperature sensitivity of an enzyme) of an enzyme-catalyzed reaction and down-regulate the enzymatic sensitivity to temperature changes (in terms of Q10), resulting in enhanced activities of b-glucosidase and arysulfatase (Paz-Ferreiro et al., 2015). A reduced Ea under biochar application indicates higher substrate affinity (in term of apparent Km) and less temperature sensitivity of the enzyme, possibly due to sorption of substrates and enzymes on biochar (PazFerreiro et al., 2012). On the other hand, soil enzymes respond quickly to soil management (e.g., organic matter amendment)

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(Bandick and Dick, 1999), and thus changes in soil properties with biochar application can reasonably influence soil enzyme activities. For the third mechanism, inhibitors released from biochar can interfere with enzyme-catalyzed reactions as well: for example, during pyrolysis, plant biochars can release a number of benzofurans, polycyclic aromatic hydrocarbons, and heterocyclic compounds, which are inhibitory compounds to soil enzymes (Lammirato et al., 2011).

that can act as microbial inhibitors (Spokas et al., 2011) may be the reason for the uncertainty in biochar-triggered pathogen resistance. To better reveal the mechanisms, the identification and quantification of reactive compounds from biochar, which are able to interfere with microbial communication, is an important research field.

2.6. Biochar affects intra- and interspecific communication of organisms

Biochar as a soil ameliorant can decrease the toxicity of soil contaminants to soil microbes (Koltowski et al., 2017). Willow biochar (pyrolyzed at 700  C) can reduce the microbe mortality while increasing the reproduction of Folsomia candida in soils contaminated with heavy metals and organic pollutants (PAHs) and can decrease the leachate toxicity to the bacterium Vibrio fischeri (Koltowski et al., 2017). The immobilization of soil contaminants (including heavy metals such as Al, Cd, Co, Cr, Mn, and Ni, and organic pollutants such as PAHs) on biochar and the consequent reduction of their bioavailability can be the main reason for the alleviated toxicity of soil contaminants to microbes and the elevated microbial biomass (Qian et al., 2013, 2015; Seneviratne et al., 2017; Zielinska and Oleszczuk, 2016). Rice straw biochar application (5% application rate) can result in an increase of the organic-bound fraction of heavy metals (Cd, Cu, Pb, and Zn) by up to 68% (Lu et al., 2017). The reduced heavy metal stress on N-fixing bacteria (Bradyrhizobium japonicum) could further benefit the N supply for plant growth (Seneviratne et al., 2017). In addition to the effect of reducing the toxicity of contaminants to soil microbes, thus enhancing microbial activity, interactions between the biochar and soil microbes have more profound effects on the environmental fate of soil contaminants, including their immobilization and degradation, which are to be discussed in a later chapter of this review.

Biochar modifies microbial cell-to-cell communication via the sorption of signaling molecules (e.g., an N-acyl-homoserine lactone, AHL) (Masiello et al., 2013) and, perhaps more importantly, by enhancing the hydrolysis of AHL (Gao et al., 2016). The intercellular signaling molecule N-3-oxo-dodecanoyl-L-homoserine lactone is an AHL that is used by many gram-negative soil bacteria (e.g., nitrogen-fixing plant symbionts and pathogens that cause soft rot in plants) to regulate gene expression and intraspecific communication (Masiello et al., 2013). When sorption is the main mechanism, this effect is dependent on pyrolysis temperature, which influences the biochar adsorption capacity (Masiello et al., 2013). High-temperature biochar (700  C), with higher specific surface area (SSA), adsorbed the signaling molecule (AHL) and interrupted the inter-cell transfer of the signal to a larger extent than moderate temperature biochar (300  C), with a much lower SSA, did (Masiello et al., 2013). By adjusting the soil pH, biochar is able to inhibit or promote the exchange of signaling compounds that regulate specific soil microbial activity (Warnock et al., 2007). Additionally, the increase in soil pH induced by biochar can drive the hydrolysis of the signaling AHLs, thereby reducing the number of bioavailable AHLs and consequently inactivating the bacterial cell-to-cell communication (Gao et al., 2016). While some signaling molecules for bacterial communication (such as AHL) are pHsensitive, some fungal signals (such as farnesol, a fungal autoinducer) are less sensitive to pH, possibly leading to shifts in the fungi to bacteria ratio (and a change in the soil microbial community structure) with biochar application (Gao et al., 2016). Biochar can modify microbe-to-plant communication in the rhizosphere (Harel et al., 2012), affect the competition between beneficial soil microbes and soilborne pathogens (Akhter et al., 2015), and trigger systemic plant defenses to soilborne pathogens (Akhter et al., 2015; Elad et al., 2011). Application of biochar successfully induced resistance against two foliar fungal pathogens (Botrytis cinerea and Leveillula taurica) for both pepper and tomato plants (Elad et al., 2010). This action is regulated by either the enhanced interaction between plant roots and the plant-growthpromoting rhizobacteria (PGPR) or fungi (PGPF) (Harel et al., 2012) or by the direct toxic effects of certain biochar components on microbial pathogens (Graber et al., 2010). Both wood biochar and green waste biochar, when amended to the potting medium, stimulated a range of general defense pathways of strawberry plants and as a result suppressed anthracnose disease caused by the fungal pathogens Botrytis cinerea, Colletotrichum acutatum, and Podosphaera aphanis (Harel et al., 2012). It was found that higher application rates (3%) of biochar induced higher defense ability against the pathogen-caused plant disease than low application rates (1%) of biochar did (Harel et al., 2012). Moreover, there are also various influences of different biochar types (wood and green waste biochar) on the competition between arbuscular mycorrhizal fungi (AMF) and the soilborne pathogens (e.g., Fusarium oxysporum f. sp. lycopersici) in the mycorrhization on tomato roots (Akhter et al., 2015). The variability in the amount and types of biochar VOCs

2.7. Biochar reduces the toxicity of contaminants to soil microbes

3. Soil carbon storage under biochar-microbe interaction Biochar can enhance soil carbon sequestration by various mechanisms. In addition to the stable carbon structure and its own recalcitrant nature (Guo and Chen, 2014; Xiao et al., 2014), biochar can affect carbon cycle dynamics that involve microbial participation, leading to either positive or negative priming effects on native organic matter, which represent enhanced or inhibited degradation of native soil organic carbon (SOC), respectively, when extra carbon sources are added (Bamminger et al., 2014; Cely et al., 2014; Kuzyakov et al., 2000; Zimmerman et al., 2011). The interactions between biochar, soil microbes, and native SOC account for the stability of biochar C and native SOC, in addition to having an overall effect on carbon sequestration (Fig. 6). The balance between the decomposition and sequestration of biochar C results in an overall contribution of biochar to the recalcitrant C pool. In addition, the effects of biochar on the microbial decomposition and physical protection of native SOC determine the fate of native SOC. The stability of biochar C, effects of biochar on the stability of native SOC, and the possible mechanisms accounting for C sequestration are the focuses of this chapter. 3.1. Carbon stability of biochar As biochars are recalcitrant for microbial degradation, they are considered stable carbon pools and have carbon sequestration potential. However, biochars produced from different conditions may have different stability, which can be reflected by the amount of aromatic structure in the biochar, as indicated by the H/C and O/C atomic ratios, and the amount of volatile matter which can be

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3.2. Effects on native organic carbon stability

Fig. 6. Schematics showing the interactions between biochar, soil microbes and native SOC, in particular their effects on soil carbon stability and the decomposition of biochar C and native SOC. The volatile matter (VM) on biochar can be used by microbes as a C source and contributes to microbial respiration. Some soil microbes can utilize aromatic C structures as well. Native SOC provides soil microbes with a major C source, and the resulting microbial decomposition of native SOC contributes to microbial respiration. On the other hand, biochar and soil microbes can protect native SOC from decomposition by forming soil aggregates with certain functional groups on the biochar and the microbial secretion of glomalin, respectively. The resulting soil aggregates of organic matter, together with the aromatic C structure of biochar as a recalcitrant C pool in soil, can contribute to the C sequestration.

utilized by soil microbes as a carbon source (Fig. 6). The reaction of soil microbes to biochar and hence the stability of biochar are driven by the feedstock types and pyrolysis temperatures of the biochars (Ameloot et al., 2015; Steinbeiss et al., 2009). Maize stover biochar had higher C, N, and P content as well as higher carbon stability than rice and wheat straw biochars did; as indicated by its aromatic C¼C structure, it also had less carbon mineralization during one year (Purakayastha et al., 2015). Oilseeds (castor bean seeds) contain high amounts of aliphatic components, and the biochar generated from them has some quickly decomposing compounds that result in greater mineralization than found in wood biochar (Rittl et al., 2015). On the other hand, pyrolysis temperatures can result in variable carbon structures of biochar, generally leading to enhanced carbon stability with increased pyrolysis temperature (Xiao et al., 2014). Biochar stability affects the microbial decomposition of biochar, which is defined as biological aging (Czimczik and Masiello, 2007; Zimmerman, 2010). It is suggested that labile and semi-labile fractions make up the non-stable biochar fractions, and the characteristics of stable and non-stable fractions of biochars determine their resistance to chemical and biological degradation (Masek et al., 2013). Microbes involved in the transformation of aromatic compounds may be the first colonizers and decomposers of fresh pyrogenic carbon, including biochar (Zimmermann et al., 2012); as such, these microbes influence the carbon stability. For example, biochar application has caused increases of the relative abundance of chitin and cellulose degraders (Chitinophaga and Cellvibrio, respectively) and aromatic compound degraders (Hydrogenophaga and Dechloromonas) on the genus level, likely being induced by the enrichment of aromatic compounds such as phenol, methylphenol, and dihydroxybenzenes in biochar (Kolton et al., 2011). Enrichment of the chitin-, cellulose-, and aromatic compound-decomposing microbial genera indicated their important role in the biochar biological aging process.

There is abundant evidence that biochar application can affect the stability of native soil organic carbon (SOC) in terms of a “priming effect” (Ameloot et al., 2014; Jien et al., 2015; Zimmerman et al., 2011). This effect can be caused by microbial community changes that are induced by biochar, which in turn increase or decrease the decomposition of native SOC (thus determining the stability of the native SOC). The effect may also be caused by the physical protection of native SOC on biochar (Fig. 6) (Ameloot et al., 2013). An alteration of the soil microbial community to a higher fungi-to-bacteria ratio can suppress the degradation of native SOC and cause a large negative priming effect up to 68% (Bamminger et al., 2014). The priming effects of biochars on native SOC are dependent on the feedstock characteristics and pyrolysis conditions, which determine the difficulty of utilizing biochar organic carbon fractions as a carbon source by soil microbes. The dominant properties of biochar that can influence the priming effects on native SOC include the carbon content, carbon aromaticity, volatile matter content, the amount of fixed and easily oxidized carbon components, and various surface properties (Cely et al., 2014). Due to the differences in carbon stability, woodchip biochar had a negative priming effect, while biochars generated from wheat husk and sewage sludge caused a positive priming effect in short-term (45 days) incubations (Cely et al., 2014). In contrast to the role of the aromatic carbon structures of biochar in the priming effect, the volatile matter content of biochar (i.e., labile carbon fraction) correlated positively with the initial biological oxygen demand (BOD), implying that positive priming effects can be due to the microbial utilization of the labile carbon fraction within the applied substrate, which can enhance the microbial growth that controls the positive priming effect on native SOC (Ronsse et al., 2013). On the other hand, biochar application can enhance the effectiveness of microbial energy utilization by simultaneously decreasing the metabolic quotient (qCO2) and increasing the soil microbial biomass (or in some cases, by reducing soil respiration with unchanged microbial biomass) (Bamminger et al., 2014; Zheng et al., 2016), which accounts for one mechanism of the negative priming effect of biochar on native SOC. The decreased qCO2 of soil microbes can be achieved by modifying the thermodynamic parameters of soil enzyme activities with biochar applications (Paz-Ferreiro et al., 2015). The oxidation of biochar surfaces during the aging process introduces oxygenated functional groups (e.g., carboxylic and ester groups) on biochar surfaces (Novak et al., 2009; Qian and Chen, 2014); therefore, it can enhance the stability of native SOC by physically protecting soil aggregates (Fig. 6) (Heitkoetter and Marschner, 2015). The introduction of the oxygenated functional groups on biochar surfaces can lead to enhanced biochar and soil CECs (Hale et al., 2011; Lin et al., 2012), and higher sorption of soil DOM on biochar surfaces can provide acidic functional groups and further increase the surface charge of the biochar-SOM aggregates (Fig. 6) (Heitkoetter and Marschner, 2015). Moreover, the oxidation-introduced carboxylic functional groups on aged biochar surfaces serve as additional binding sites, which can intimately attach mineral phases composed of Al, Si, and Fe and form large aggregates (Lin et al., 2012; Qian and Chen, 2014). As a consequence, biochar is considerably involved in the preservation of the soil humus, including humic acids (HAs), fulvic acids (FAs) and humins (HMs) in soil, which can be ascribed to the presence of abundant oxygenated functional groups that favor soil aggregation formation via sorption (Hua et al., 2015). To evaluate the effects of biochar application on the soil carbon sequestration and the deposition of native SOC, it is crucial to distinguish between the microbial decomposition of organic

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carbon from biochar and that from native SOC. For this purpose, a two-component model can be used to indicate the carbon stability of biochar carbon and native SOC in a system receiving biochar application (Galvez et al., 2012; Murray et al., 2015; Zimmerman et al., 2011), as shown in equation (1): Cmineralized (%) ¼ f  (1
e

-k1t

) þ (100
f)  (1
e

-k2t

) Eq.

(1)

where Cmineralized is the carbon fraction that is cumulatively mineralized, f is the carbon fraction (%) of the active or fast turnover carbon pool (native SOC), k1 and k2 represent the mineralization rate constants of the active and resistant carbon pools (biochar carbon), respectively, and t is the incubation time in days. The mean retention time (MRT) of the respective carbon pools can be calculated as the inverse of the mineralization rate constants k1 and k2, i.e., 1/k1 and 1/k2 (Murray et al., 2015). After a given application of biochar, the shifts of the parameters within this model can indicate changes in carbon stability and the effectiveness of carbon sequestration. This model indicates that less than 3% of the applied biochar was lost through CO2 evolution, with an MRT of 600 years at an annual mean temperature of 26  C (or, 3264 years at 10  C) (Major et al., 2010). The model also indicates that biochar can reduce the mineralization rate constant of native SOC and increase its MRT, especially in acidic soil (Galvez et al., 2012).

3.3. Mechanisms of biochar-enhanced soil carbon sequestration The factors involved in the maintenance of soil carbon stability with biochar application (Fig. 7) include: (1) the recalcitrant nature of biochars, due to their aromatic carbon structure and the crystal silicon structure in silica-carbon complexes formed during pyrolysis (Guo and Chen, 2014; Xiao et al., 2014); (2) the interaction between surface oxygenated functional groups such as carboxyl with soil minerals are favorable to maintaining the carbon stability of biochar during the aging process (Chen et al., 2015; Li et al., 2014); (3) the adsorption of soil organic matter that is beneficial for soil aggregate formation (George et al., 2012); (4) the availability

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of soil arbuscular mycorrhizal fungi that are able to secrete glomalin, the protein that promotes soil microaggregation due to its impact on hydration (King, 2011); and (5) the influences on the soil microbial community and the modification of soil enzyme activities that control soil organic carbon decomposition (Paz-Ferreiro et al., 2015). Among these factors, the protection of soil aggregates and modification of enzyme activities are related to microbial activity and are to be mainly discussed here. Biochar can act as a binding agent of soil organic matter during the formation of aggregates and can thus promote the macroaggregation (George et al., 2012) that enhances the stability of soil aggregates (Sun and Lu, 2014). The number of stable macroaggregates (>250 mm) can be increased by approximately 25% in the case of a sandy soil after 80 days of incubation, which is helpful for soil pore water retention, soil carbon sequestration, and microbial protection (Awad et al., 2013). The aggregation of large-size soil particles is partly due to the adsorption of soil organic compounds, such as humic acids, on biochar surfaces (Kasozi et al., 2010). Changed patterns of soil aggregation will alter microbial community structures in soil, since gram-positive and gram-negative bacteria tend to possess different strategies in occupying soil particle fractions. There is a greater abundance of gram-positive bacteria in silt and clay fractions, which are older and have more microbially processed soil organic matter, while there are more gram-negative bacteria in fractions with larger (>200 mm) aggregates, as these bacteria are fond of fresh plants as C sources (Zhang et al., 2015). Similarly, fungal abundance was found to be reduced with the decreasing size of soil aggregation due to the C substrates being less available in soil fractions with aggregates smaller than 200 mm (Zhang et al., 2015). The effects of biochar amendment on soil microaggregation are not only related to the direct adsorption of soil organic matter on biochar surfaces but also to the alteration of the microbial community (especially fungi) that produces proteins such as glomalin, hydrophobins, and chaplins that positively influence microaggregate formation and stability (King, 2011; Rillig et al., 2005). Glomalin, produced by a phylogenetically narrow range of arbuscular mycorrhizal fungi within the order Glomales, has the function to increase the hydrophobicity and stability of microaggregates (Rillig et al., 2002). Hydrophobins, occurring widely in numerous basidiomycetes and ascomycetes, take vital parts in mycelia formation and microaggregate protection (Kershaw and Talbot, 1998). Chaplins, released mainly by streptomycetes in phylum Actinobacteria, are proteins of bacterial origin that generate hydrophobic soil surfaces (Gebbink et al., 2005; Sawyer et al., 2011). The hydrophobic nature of the above proteins prevents water penetration into soil organic matter and results in microaggregate formation (Rillig, 2005). Thus, biochar-enhanced soil carbon sequestration can be reasonably attributed to biochar functions on soil microbiota that maintain soil microaggregate stability through the protein secretion. 4. Fate of soil contaminants under biochar-microbe interactions 4.1. Immobilization of contaminants by biochars

Fig. 7. Proposed carbon sequestration mechanisms affected by biochar, including: (A) the stable carbon structure of biochar (aromatic carbon structure and crystal silicon structure in silica-carbon complexes) (Guo and Chen, 2014; Xiao et al., 2014); (B) the reaction of biochar with soil minerals, which forms a complex structure that protects biochar from microbial degradation (Chen et al., 2015; Li et al., 2014); (C) biochar adsorption of soil organic matter (SOM), forming aggregates that protect the degradation of SOM (George et al., 2012); (D) protection of soil aggregates by fungal hyphae and the secreted glomalin (King, 2011); (E) modification of soil enzyme activities that control soil organic carbon decomposition (Paz-Ferreiro et al., 2015). These processes can reduce the CO2 emission from the decomposition of biochar and SOM.

Biochar is an effective sorbent that immobilizes organic contaminants and heavy metals through various mechanisms (Ahmad et al., 2014; Chen and Chen, 2009; Chen et al., 2008a, 2008b, 2012b, 2012c). Electrostatic attraction, polar and non-polar organicattraction to the carbonized phase of biochar, and partitioning to the non-carbonized phase of biochar are forces involved in the interactions of biochar with organic contaminants (Chen and Chen, 2009; Chen et al., 2008a; Huang and Chen, 2010). Ion exchange,

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anionic metal attraction, precipitation, and cationic metal attraction are efficient mechanisms for the interaction of biochar with inorganic contaminants (Ahmad et al., 2014, 2016; Cao et al., 2009; Qian and Chen, 2013; Xu and Chen, 2015). In some cases, biochar can be even more effective than activated carbon to immobilize heavy metals, e.g., a study found six times more effective Pb sorption (up to 680 mmol Pb kg
1) by dairy manure-derived biochar than activated carbon, mainly attributed to Pb precipitation (84e87%) in form of bePb9(PO4)6 and Pb3(CO3)2 as well as surface sorption (13e16%) via C¼C (p-electron) and eOePb bonds through interactions with biochar (Cao et al., 2009). The sorption of certain herbicides (such as simazine) in micropores or on the surfaces of biochar prevents their accessibility to microbial cells, extracellular enzymes, and plants, in addition to preventing them from leaching into the groundwater (Jones et al., 2011). A simple modification of biochar with iron oxide can endow such magnetic biochar with a hybrid sorption capability of organic pollutants and phosphate, which enables biochar a multifunctional material for agricultural and environmental applications (Chen et al., 2011). Recent research reveal the novel roles of biochar on alleviating the aluminum (Al) toxicity (Qian and Chen, 2013) by modifying the speciation of Al(OH)2þ or Al(OH)3þ, rather than by a direct electrostatic attraction of Al3þ with negatively charged sites on biochar (Qian et al., 2013). The oxidation of biochar surfaces during the aging process can provide extra binding sites for heavy metals Cd and Al, though the competition sorption between Cd and Al exist (Qian and Chen, 2014; Qian et al., 2015). Novel roles of biogenic silicon in biochars are found to alleviate Al phytotoxicity and to prevent plant uptake of Al, indicating that Si particles can reduce the amount of soil exchangeable Al and that Si released from biochars can form Si-Al compounds in the epidermis of wheat roots (Qian et al., 2016). The above research show the necessity to further investigate the elemental composition of biochar and the role of biochar on the biogeochemical cycles of soil elements, which can be involved with soil microbes. 4.2. Influences on contaminant transformation and dissipation Evidence has accumulated that biochars are able to act as electron shuttles between microorganisms and pollutants to enhance microbial degradation (Fig. 8) (Yu et al., 2015). Biochar can bridge the electron transfer from the microbial cells to the mineral (Fig. 8) (Kappler et al., 2014), acting as the direct electron acceptor of acetate for microbial extracellular respiration and growth (Yu et al., 2015) and as the electron donator to stimulate the microbial reduction of the Fe(III) oxyhydroxide mineral ferrihydrite (Fig. 8) (Kappler et al., 2014). When the reduction occurs as a result of the electron transfer from Geobacter sulfurreducens, biochar may further serve as an electron donator for the metabolism of other microorganisms, in such way to catalyze biochemical processes (Kappler et al., 2014; Yu et al., 2015). The presence of biochar has significantly enhanced the degradation of pentachlorophenol (PCP) by a species of bacteria (Geobacter Sulfurreducens), which is hardly able to degrade PCP alone; this enhancement was ascribed to biochar-facilitated electron transfer between microbial cells and PCP molecule (Yu et al., 2015). The electron transfer capabilities of biochar are influenced by the pyrolysis temperature; the effects are mainly attributed to the phenolic moieties formed under low temperatures and quinonyl groups formed under moderate and high pyrolysis temperatures (Fig. 8) (Kluepfel et al., 2014). The pool of redox-active moieties (RAMs) in biochar is dominated by different components and is quantified by its electron exchange capability (EEC), which is mainly composed of the electron-donating capability (EDC) of the electron donators, e.g., phenolic moieties in biochars, and the

electron-accepting capability (EAC) of the electron acceptors, e.g., quinonyl moieties (Kluepfel et al., 2014; Yang et al., 2016a). Biochar could store hundreds of micromoles of electrons per gram of mass, depending on the feedstock and hydrolysis temperatures during the process of accepting electrons, and could further donate them to minerals, causing their reduction (Kappler et al., 2014). It is suggested that the redox activity of biochar was related to the water-soluble organic components in biochar suspensions and dissolved organic matter (DOM), which is composed of phenols (Graber et al., 2014). In contrast to the dissolved compounds that can be extracted with water from biochar particles, the solid phase of biochar particles was recognized to be responsible for the electron transfer that caused the microbial ferrihydrite reduction (Kappler et al., 2014). Direct contact between microbial cells, pollutant molecules, and biochar particles through surface adsorption can favor the electron transfer between them and further accelerate the pollutant degradation, which is a process that needs further investigation. Changes in the redox activity of biochar are able to enhance the direct interspecific electron transfer between microbial cells (DIET) and direct extracellular electron transfer between microbial cells with minerals, organic matter, contaminants, and heavy metals (DEET), resulting in the degradation of organic pollutants (e.g., pentachlorophenol) and the decomposition of organic matter (e.g., humic acids) (Fig. 8). Together with RAMs, the electrical conductivity (EC) of biochar contributes to its role as an electron shuttle between microbial cells and pollutants, with a proposed mechanism that involves the formation of a p-p network within the graphite-like aromatic structure of biochar, which has a high EC. The shuttle function of biochar is responsible for the electron transfer to the organic compounds (e.g., PCP) that are adsorbed on its surface (Chen et al., 2014; Yu et al., 2015), as shown in Fig. 8. Biochar had various ECs, depending on feedstock type and pyrolysis temperatures, roughly ranging from 2.11 to 4.41 mS cm
1 g
1 in pine biochar and much higher (from 216 to 2217 mS cm
1) in manure biochar (Cantrell et al., 2012; Chen et al., 2014). In addition to the biochar-facilitated microbial degradation of hazardous pollutants, biochar can reduce the adverse effects from organic contaminants and heavy metals on soil microbes and plants by adsorbing such hazards and further directing electron transfer between them (Choppala et al., 2012). For metal(loid) ions, such as Cr and As, whose toxicity depends on their valence, the regulation of electron transfer by biochar may control their fate and risks (Fig. 8). Biochar DOM serves both as electron donor and acceptor to reduce toxic Cr(VI) to the less toxic and relatively immobile Cr(III) and to oxidize As(III) to As(V). This function is most likely due to the high reactivity of Cr with many functional groups in biochar DOM that are able to donate electrons to reduce Cr(VI) and oxidize As(III) (Choppala et al., 2012; Dong et al., 2014). Losses of Cr(VI) from suspensions of biochar-applied soil and decreased toxicity, induced by Cr, to sunflowers has been found as well (Choppala et al., 2012). By immobilizing Cr and reducing its concentration in soil pore water, biochar application can mitigate Cr contamination and enhance soil microbial activity and sunflower plant growth (Choppala et al., 2012). The activation of persistent free radicals formed in biochar (Fig. 5) is commonly considered to be the mechanism of biocharfacilitated contaminant degradation (Fang et al., 2014a, 2015a, 2015b). The oxidation of biochar during the aging process releases redox-active, acidic and phenolic organics, and the interaction with transition metals produces PFRs, making biochar continuously participate in redox reactions and hazardous organic compound degradation (Fang et al., 2014a; Graber et al., 2014). Free radicals in biochar exhibited excellent reactivity to generate OH with H2O2 decomposition, consuming approximately 12 spins of

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Fig. 8. Schematic diagram showing mechanisms of biochar acting as an electron shuttle enhancing direct interspecific electron transfer between microbial cells (DIET) and direct extracellular electron transfer between microbial cells with minerals, organic matter, contaminants, and heavy metals (DEET) (Chen et al., 2014; Dong et al., 2014; Kluepfel et al., 2014; Xu et al., 2016; Yu et al., 2015). Electrons can be transferred from one microbial cell to the component with an electron-accepting capacity (EAC, e.g., quinone) in biochar, shuttled between EAC and the component with an electron-donating capacity (EDC, e.g., phenol) and be accepted either by another microbial cell, the organic contaminant (pentachlorophenol), the hematite minerals, or humic acids, enhancing their transformation or degradation. The graphite-like structure on biochar can realize the conductivity. Electron transfer occurring in biochar dissolved organic matter (DOM) facilitates the transformation of metal(loid)s to nontoxic forms: Cr(VI) to Cr(III) and As(III) to As(V).

free radicals to trap one [OH] molecule (Fang et al., 2015b); sulfate radicals (SO
4 ), formed by persulfate activation, could efficiently degrade contaminants including 2-chlorobiphenyl (2-CB), polychlorinated biphenyls (PCBs), and diethyl phthalate (DEP) (Fang et al., 2014a, 2015a, 2015b). As a practical application, the combination of biochar and microbial remediation strategies is proposed as an effective method for soil organic contaminant remediation (e.g., for soil PAH pollution management) (Chen and Ding, 2012; Chen et al., 2012a; Garcia-Delgado et al., 2015). Using biochar as carriers of PAHdegrading bacteria (Pseudomonas putida, Pseudomonas aeruginosa, Acinetobacter radioresistens, and an unidentified indigenous bacterium), the immobilized-microorganism technique (IMT) finds its effective application on alleviating soil PAHs contamination (Chen et al., 2012a; Galitskaya et al., 2016). For the strategy of combined biochar-microbial remediation, biochars are used as a pretreatment to immobilize and concentrate organic contaminant PAHs, followed by an inoculation of PAH-degrading microbes (e.g., white rot fungus Phanerochaete chrysosporium and Pleurotus Ostreatus) to perform the final degradation of the PAHs in soil (Chen and Ding, 2012; Garcia-Delgado et al., 2015). The sequential application of biochar and white rot fungi has induced higher degradation of phenanP P threne, anthracene, fluoranthene, pyrene, 3-rings, 4-rings, and P 13 PAH than the single addition of this species of fungi (Chen and Ding, 2012; Garcia-Delgado et al., 2015). Two possible mechanisms that could be involved in the influence of biochar on the microbial dissipation of contaminants: (1) the attachment of microbial cells to biochar particles and biochar-enhanced soil microaggregates that provide suitable habitats and prevent harsh environmental changes (such as pH changes, water retention, temperature

changes and predators) to ensure microbial growth that contributes to contaminant degradation; (2) persistent free radicals could assist biochar as a catalyst and as an electron shuttle to enhance the electron transfer between the microbial cells and pollutant molecules, thus facilitating heavy metal transformation and organic pollutant degradation (Kappler et al., 2014; Yu et al., 2015). 5. Perspective and conclusions Heterogeneous effects on soil remediation are found from the complex combination of biochar and soils with various properties. Through the massive research with respect to biochar effects on soil physiochemical characters and microbial activities, the most influential factors point to feedstock types and pyrolysis temperature, which determine biochar properties including surface properties, structures, elemental composition, redox capacity, conductivity, pH, CEC, and VOCs. Such properties play key roles in microbial activity and soil process. According to the specific purposes, different biochar types should be considered. To improve soil pH, especially for acidic soils, the utilization of wood biochar produced under high temperature can be proper as it loses most of the acidic functional groups during pyrolysis. To enhance soil fertility, crop residue biochar and manure biochar with high pyrolysis temperatures have the priority, because their high ash content can improve soil CEC, facilitating nutrient maintenance for microbial and plant growth. To enhance carbon sequestration, high temperature biochar, especially from woods, should be considered because they generally have stable carbon structure and high C/N ratio that make them recalcitrant to microbial decomposition. To control soilborne pathogens, low temperature biochars may have the

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advantages, because the relatively high small molecular VOCs on such biochar can be microbial inhibitors. Moreover, moderate temperature biochar with relatively higher adsorption capacity and electron transfer capability can be considered for the purpose of soil contamination remediation. Understanding the key roles of biochar properties on microbial activity in different soil types is essential to answering the most challenging questions, such as: (1) Under which conditions can the application of biochar reach the desired benefits, and how can the trade-offs between various environmental effects of biochar be balanced? (2) Why, mechanistically, is a given biochar beneficial in some soils but not in other soils? (3) How can the biological effects of biochar be better predicted? Focused on certain key functions of biochar, this review synthesizes the biochar-microbe interaction mechanisms pertinent to those questions, covering: (1) the microbial community modification by biochar, via alteration of nutrient availability and soil characters, (2) the electron transfer between microbial cells and contaminants facilitated by biochar, (3) the potential microbial toxicity triggered by biochar, (4) the carbon stability in biochar with regard to the role of microbial degradation and the effects of biochar on the soil organic carbon stability, and (5) the mechanisms of biochar-modified plant resistance to soilborne pathogens. New developments in the metagenomics analysis of soil microbe genes will no doubt be pivotal to uncovering the hidden dimensions of biochar-microbe interactions that link biochar properties with microbial functions. Understanding the mechanisms of the interaction between biochar and soil microbes is essential to revealing the mechanisms of the heterogeneous effects of biochar on soil remediation. Competing financial interests The authors declare that they have no competing interests. Acknowledgements This project was supported by the National Natural Science Foundation of China (Grant 21425730, 21537005, 21621005, and 21607125), The National Key Technology R&D Program (2015BAC02B01), and the Postdoctoral Science Foundation of China (Grant 2015M581943). Special thanks give to the anonymous reviewers who contribute greatly to improve the quality of this review. Appendix ASupplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2017.04.032. References Abel, S., Peters, A., Trinks, S., Schonsky, H., Facklam, M., Wessolek, G., 2013. Impact of biochar and hydrochar addition on water retention and water repellency of sandy soil. Geoderma 202, 183e191. Abit, S.M., Bolster, C.H., Cai, P., Walker, S.L., 2012. Influence of feedstock and pyrolysis temperature of biochar amendments on transport of Escherichia coli in saturated and unsaturated soil. Environ. Sci. Technol. 46, 8097e8105. Ahmad, M., Ok, Y.S., Kim, B.-Y., Ahn, J.-H., Lee, Y.H., Zhang, M., Moon, D.H., AlWabel, M.I., Lee, S.S., 2016. Impact of soybean stover- and pine needle-derived biochars on Pb and As mobility, microbial community, and carbon stability in a contaminated agricultural soil. J. Environ. Manag. 166, 131e139. 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, 19e33. Akhter, A., Hage-Ahmed, K., Soja, G., Steinkellner, S., 2015. Compost and biochar alter mycorrhization, tomato root exudation, and development of Fusarium oxysporum f. sp lycopersici. Front. Plant. Sci. http://dx.doi.org/10.3389/ fpls.2015.00529. Ameloot, N., Graber, E.R., Verheijen, F.G.A., De Neve, S., 2013. Interactions between

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