Biochar-Facilitated Soil Remediation: Mechanisms and Efficacy Variations

Recent research suggests that biochar amendment is a promising approach to mitigate soil contamination via immobilizing heavy metals and organic pollutants. Through intensive literature review, this paper was aimed to better understand the processes, mechanisms, and effectiveness of biochar in immobilizing chemical contaminants in soil. The quality characteristics of biochar as a soil amendment varied greatly with the feedstock materials and the pyrolysis conditions. Biochar products from different sources demonstrated remarkably diversified capacities and efficiencies for stabilizing soil contaminants. Soil-incorporated biochar was able to stabilize Cd, Cu, Ni, Pb, and Zn and reduce their bioavailability through enhanced sorption (based on electrostatic attraction, ion exchange, and surface complexation) and chemical precipitation (incurred from soil pH elevation and ash addition of carbonates and phosphates). The stabilization efficacy was largely determined by cation exchange capacity, pH, and ash content of the biochar. Biochar amendment increased the mobility of anionic toxic elements [e.g., CrO42-, AsO3-, and Sb(OH)6-] in soil. Soil-incorporated biochar was also able to adsorb non-polar organic compounds (through pore filling, partition, and hydrophobic effect) and polar organic compounds (via H-bonding, electrostatic attraction, specific interaction, and surface precipitation). The adsorption efficiency was controlled by the biochar surface properties specific surface area, microporosity, and hydrophobicity. Biochar may facilitate the mineralization of organic pollutants by enhancing soil microbial activities. The effectiveness of biochar-facilitated soil remediation was case specific, changing with the biochar source, amendment rate, placement, soil type, and pollutant species. More field studies are needed to evaluate the long-term effectiveness of biochar-facilitated soil remediation under practical circumstances.


INTRODUCTION
Biochar is charcoal prepared by pyrolytic processing (i.e., O 2 -absent heating at 300-700 • C) of residual biomass materials and used as a soil amendment in agricultural and environmental applications (Lehmann and Joseph, 2012). The black carbonaceous solid is porous, environmentally recalcitrant, and abundant in surface functional groups . The relatively high porosity and surface functionality engender biochar with great specific surface area (SSA) and cation exchange capacity (CEC), enabling the material to retain water, nutrients, and pollutants in soil (Ahmad et al., 2014). In addition, biochar contains significant portions of labile organic carbon (OC) and possibly plant nutrients (e.g., N, P, K, Ca, Mg, and S) (Song and Guo, 2012). Intensive laboratory studies and field trials have demonstrated that appropriate biochar amendment is effective to ameliorate soil physical, chemical, and biological properties, increase crop productivity, and reduce the bioavailability of heavy metals and organic contaminants in soil (Chan et al., 2007;Park et al., 2011;Novak et al., 2016;Ghorbani et al., 2019). As a promising soil amendment, biochar has been explored in uses of soil health improvement, pollution alleviation, land reclamation, and climate change mitigation.
Soil contamination by heavy metal(loid)s (e.g., Cd, Cr, Hg, Pb, Cu, Zn, As, Co, Ni, and Se) and persistent organic pollutants [POPs, such as polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), polychlorinated dibenzoρ-dioxin (PCDD), and polychlorinated dibenzofurans] is a worldwide problem that threatens environmental sustainability, food safety, and human health (Fabietti et al., 2009;Su et al., 2014;Sun et al., 2018). To control the hazardous effects and ideally restore the ecosystem services of contaminated soils, an array of in situ and ex situ remediation techniques have been developed, including surface capping, encapsulation, landfilling, soil flushing, soil washing, soil venting (air sparging), vacuum ventilation (vapor extraction), thermal extraction, electrokinetic extraction, chemical degradation, stabilization, solidification, vitrification, bioremediation, and phytoremediation (Zhu et al., 2010;Liu L. et al., 2018). These techniques employ physical, chemical, thermal, electrical, and biological methods and processes to contain, immobilize, and ultimately eliminate soil contaminants. In field practice, all the available techniques demonstrate particular advantages and disadvantages of remediation efficiency, cost effectiveness, and applicability (Liu L. et al., 2018).
Recently, biochar amendment has been investigated to alleviate soil contamination and facilitate soil remediation (Chai et al., 2012;Houben et al., 2013;Koltowski et al., 2016;Ippolito et al., 2017). Existing literature reviews suggest that biochar amendment immobilizes heavy metals and POPs in contaminated soils and reduces their bioavailability primarily through precipitation, electrostatic interaction, surface adsorption, structural sequestration, and facilitated decomposition; the decontamination efficacy varies with the biochar source, amendment rate, soil type, and pollutant species (Ahmad et al., 2014;O'Connor et al., 2018;Zama et al., 2018;Kumpiene et al., 2019;Yuan et al., 2019). Most of the reported research trials were in short-term laboratory experiments. The long-term (e.g., >5 years) effects of biochar amendment on contaminant mobility and bioactivity in field soils remain unclear. By comprehensively reviewing the literature, this paper was aimed to further understand the mechanisms and the processes through which biochar amendment mitigates soil contamination, with the ultimate goal to provide instructions of appropriate biochar utilization for effective soil remediation.

QUALITY VARIATIONS OF BIOCHAR AS A SOIL AMENDMENT
Biochar products from different sources vary widely in characteristics and function capacity as a soil amendment. Biochar is produced from biomass materials using the thermochemical technique pyrolysis, through which organic residues are heated in O 2 -free or highly limited, ambientpressure environments for a certain time to be carbonized into charcoal, with the generation of pyrolysis bio-oil and syngas as byproducts . Common biochar feedstocks extend to forest debris, crop residues, food processing waste, and manures including sewage sludge and biosolids ( Table 1). These biomass materials are significantly different in organic and ash compositions, attributing to the notable quality variations of the resulting biochar products. The carbonization (pyrolysis) conditions further influence the quality characteristics of biochar. Three parameters are usually used to manipulate the carbonization conditions: pyrolysis (peak) temperature, solid residence time, and heating rate, each stretching over a wide range of values . The pyrolysis temperature for biochar production is mostly in the range of 300-700 • C ( Table 1). A higher temperature accelerates the carbonization process, allowing the pyrolytic transformation of biomass to reach a deeper level and be completed in a shorter time (Song and Guo, 2012;Chen et al., 2017). Complete pyrolysis is critically important to transform all feedstock OC into carbonized, pyrogenic OC (i.e., altered, amorphous C structure). The solid residence time for achieving complete pyrolysis is determined by the pyrolysis temperature and the heating rate, ranging from seconds to days. The heating rate may be as low as <1 • C s −1 and as high as >200 • C s −1 , varying with the pyrolysis temperature, feed characteristics (particle size, moisture content, and density), and feed mass flow . Biochar products from incomplete pyrolysis also contain noticeable portions of uncarbonized C (i.e., with crystalline character of the precursor materials) (Chun et al., 2004;Keiluweit et al., 2010).
Biochar is porous, showing a rough morphological surface with honeycomb-like anatomical or other irregular structures under the microscope (Figure 1). The inherent micropores (diameter in the range of 0.8-235 µm and mostly <22 µm, with average at 10 µm; Hardie et al., 2014) engender the material a relatively high intrapore volume (e.g., 0.9-1 cm 3 g −1 ; Batista et al., 2018) and low envelope density (e.g., 0.2-0.6 g cm −3 ; Joseph et al., 2019). Biochar is composed primarily of amorphous, aromatic carbon, and possesses abundant O-containing surface functional groups (e.g., -C=O, -COOH, and -OH). A disorderly stacked graphene sheet structure has been proposed for biochar (Figure 1). The aromaticity of biochar generally increases while the surface functionality decreases as the pyrolysis temperature is elevated (Song and Guo, 2012;Zhao et al., 2016). This is largely a result of the progressive losses of aliphatic C-H, olefinic C=C, carbonyl, carboxyl, and hydroxyl groups at a higher pyrolysis temperature (Fu et al., 2012;Tan et al., 2015).
Biochar is the major product of slow pyrolysis and yet the byproduct of fast pyrolysis (aiming at pyrolysis bio-oil) and gasification (pyrolysis combined with partial oxidation-aiming at syngas). Within each of the three thermochemical techniques, the carbonization conditions (i.e., temperature, solid residence time, and heating rate) can be adjusted to optimize major product generation. Even with the same feedstock, biochars from gasification and fast pyrolysis are evidently lower in OC content and higher in ash content than the products from slow pyrolysis. Overall, biochars from different sources demonstrated a pH value (in 1:5 solid/water extract) ranging from 5.4 to 12.4 and mineral ash content from 1.1 to 82.0% ( Table 1). The pH is a comprehensive expression of the water-soluble base cations (e.g., Na + , K + , Ca 2+ , and Mg 2+ ) and organic acids (e.g., formic acid, acetic acid, hydroxybutyric acid, and benzoic acid) in biochar. The base cations originate from the mineral ash components in the feedstock, whereas the organic acids are generated during biomass pyrolytic transformation. Biochar products from complete pyrolysis contain minimal organic acids and demonstrate a pH value typically >7.5 to reflect the presence of basic ash minerals. In general, the pH of biochar is related to its mineral ash content, with higher-ash-content products showing a higher pH value ( Table 1). Biochars from higher-temperature pyrolysis are normally greater in ash content ( Figure 2) and pH, and the products derived from manure and rice plant residues are higher in ash content and pH than those derived from wood and other plant residues. The OC content of biochar ranged from 12.0 to 93.9%, generally decreasing as the ash content increased. For low-ash feedstocks (e.g., wood and many other plant residues), the derived biochars showed an increasing trend in OC content as the pyrolysis temperature was elevated, whereas biochars derived from ash-rich feedstocks (e.g., manures, rice plant residues, and bones) illustrated an opposite trend ( Table 1). Biochars also contain more or less the plant nutrients N, P, K, and others. The N, P, and K contents of the reported biochars are 0.5-48.0, 0.1-198.0, and 1.4-91.5 g kg −1 , respectively, being largely correlated with the products' ash contents. Most of the feedstock P and K would be recovered in biochar, yet the vast majority of feedstock N would be lost to pyrolysis vapors (biooil and syngas), especially at high pyrolysis temperatures (e.g., >500 • C) (Song and Guo, 2012). The nutrients in biochar are slowly releasable , supplementing the nutrient supply of amended soils to support plant growth. Sewage sludgeand bone-derived biochars may contain P at >100 g kg −1 and therefore serve as an ideal source of P to crops. Moreover, biochars generated from specialty biomass like municipal organic solids and contaminated-soil-grown plant materials may contain toxic elements at above-threshold concentrations for use as a soil amendment (Buss et al., 2016). Biochars demonstrate a CEC value of 5-162 cmol c kg −1 and a Brunauer-Emmett-Teller (BET)-SSA value of 0.4-448 m 2 g −1 ( Table 1). The CEC indicates the capacity of biochar to adsorb cationic nutrients and metal contaminants, while the SSA implicate biochar's ability to retain water, gases, and organic molecules. Research has revealed that biochar is generally sorptive to polar gases (e.g., H 2 S, NH 3 , NO x , and volatile organic compounds), water-soluble metal ions (e.g., Cu 2+ , Pb 2+ , Zn 2+ , and Cd 2+ ), and various organic pollutants (e.g., agrochemicals, antibiotics, and PAHs) (Ahmad et al., 2014;Tan et al., 2015). FIGURE 1 | A scanning electron micrograph of corn stover-derived biochar (left), an aromatic-C molecular biochar model (middle), and a disorderly stacked graphene sheet structure proposed for biochar (right) showing the high porosity, aromaticity, and surface functional group abundance of the material (Lehmann and Joseph, 2012;Chen et al., 2017;Zhao et al., 2017). The CEC of biochar is originated primarily from the surface functional groups, while SSA is related to biochar's porosity. A higher pyrolysis temperature in the range of 300-700 • C favors micropore development while promoting O-containing functional group elimination, resulting in biochar products with increased SSA yet reduced CEC (Figure 2). Biochars with high ash contents, like those derived from manures, typically demonstrate low SSA values, likely due to ash blockage of the inherent micropores. After applying to soil, the SSA and CEC of biochar may gradually increase over time as a result of natural weathering (Steiner, 2016).
As a soil amendment, biochar has a vital advantage over compost and other raw bio-materials in its high environmental recalcitrance. The mean residence time of biochar in the natural soil environment was estimated at 90-1,600 years (Singh et al., 2012). The environmental stability of biochar originates from its condensed aromatic carbon. Biochars with higher aromaticity and greater C condensation are generally more recalcitrant (Lehmann et al., 2015). Biochar OC can be nominally divided into three fractions: labile, intermediate stable, and stable (Singh et al., 2012). Research suggests that >65% of the OC in biochars with H/OC molar ratio <0.7 would remain in natural soils 100 years after field application (Joseph et al., 2019). Biochars prepared at higher pyrolysis temperatures generally contain a greater proportion of stable OC (Figure 2).
The International Biochar Initiative recommends the following parameters for evaluating the quality of biochar as a soil amendment: pH, lime equivalence, EC, mineral ash content, OC content, SSA, H/OC molar ratio, particle size distribution, germination inhibition assay, total and available plant nutrients, and the presence of inorganic and organic pollutants. The OC content needs to exceed 10% and the H/OC molar ratio surrender 0.7 for a pyrogenic material to be qualified as biochar (IBI, 2015). The European Biochar Certificate (EBC) program suggested similar quality variables for biochar: pH, EC, volatile matter content, total ash content, total carbon (TC) content, SSA, H/OC molar ratio, O/C molar ratio, water content, bulk density, macronutrient (N, P, K, Ca, and Mg) contents, and contents of heavy metals and organic contaminants. A certified biochar product is required to contain ≥50% TC at ≤0.4 O/C and ≤0.7 H/OC molar ratios. The contents of Zn, Pb, Cu, Cr, Ni, Cd, and Hg are capped at 400, 150, 100, 1.5, 90, 50, 1.5, and 1 mg kg −1 , respectively, and PAHs, PCBs, and PCDD/Fs at 12, 0.2, and 0.00002 mg kg −1 , respectively (EBC, 2012). Considering the mechanisms and the processes through which biochar amendments facilitate soil remediation (as discussed below), biochars selected for soil remediation should be weighted more in pH, lime equivalence, SSA, CEC, and nutrient (especially P) content when the TC content and H/OC molar ratio criteria are met.

BIOCHAR FOR REMEDIATION OF HEAVY-METAL-CONTAMINATED SOILS
Biochar has been explored for mitigating soil heavy metal contamination. Reported research using laboratory vessel, greenhouse pot, and field plot experiments suggests that biochar is capable of effectively sorbing heavy metal cations from water and immobilizing heavy metal elements in soil and therefore serves as a promising amendment for reducing the ecotoxicity of heavy-metal-contaminated soils Ahmad et al., 2014;Tan et al., 2015;O'Connor et al., 2018).

Mechanisms
Unlike soil washing, leaching, and extraction that ultimately remove heavy metals from contaminated soils (Liu L. et al., 2018), biochar amendment does not eradicate but stabilizes heavy metals in soil, transforming the toxic elements into less soluble and less bioaccessible forms. The overall process is rather similar to that of chemical stabilization in which treatment agents such as phosphate-and carbonate-containing chemicals are applied to contaminated soils to react with heavy metal contaminants and convert them into precipitates (Houben et al., 2013;Liu L. et al., 2018). The bioavailability and the ecotoxicity of heavy metals in the treated soils are consequently reduced to below the risky level, restoring the desirable soil function to grow safe crops. In biochar-facilitated soil remediation, selected biochar products are applied at appropriate rates to heavy-metal-polluted sites and thoroughly mixed with the contaminated soil. The applied biochar interacts with heavy metals in the soil, adsorbing heavy metal ions on the pore surfaces and potentially transforming them into hydroxide, carbonate, and phosphate precipitates. As the water-soluble, bioactive fraction of heavy metals in soil decreases, potential uptake and bioaccumulation of heavy metals by soil organisms (including plant roots) are minimized (Ahmad et al., 2014). Since heavy metals are not removed and the immobilization effect may diminish over time, biochar-amended soils need to be regularly monitored for heavy metal toxicity.
Biochar stabilizes cationic heavy metals primarily through sorption and chemical precipitation (Figure 3). Biochar possesses various surface functional groups such as hydroxyls, carbonyls, and carboxyls (Tan et al., 2015). Depending on the pH, dissociation or protonation of these functional groups entails biochar electrical charges. The zeta potential of canola-strawand peanut-straw-derived biochars became more negative as the solution pH increased between 3.0 and 8.0, indicating a greater negative charge density on the biochar surface at a higher pH (Xu et al., 2011). Biochars derived from corn stover, red oak, cottonseed hull, pecan shell, and pure cellulose through 200-900 • C slow pyrolysis showed pH ZNC (pH at which biochar has zero net charge) in the range of 3.1-8.5, increasing as the pyrolysis temperature was raised (Uchimiya et al., 2011a;Banik et al., 2018). This corroborates that biochars produced at higher temperatures had a lower CEC value owing to further losses of surface functional groups (Figure 2), as a higher pH ZNC indicates fewer negative charges on the surface. At pH <pH ZNC , biochar may be positively charged and therefore able to adsorb anionic pollutants such as NO − 3 , AsO − 3 , and CrO 2− 4 (Fidel et al., 2018). Both hydrolyzable non-bridging (degraded and negatively charged at high pH) and non-hydrolyzable bridging (positively charged even at high pH) oxonium groups (oxygen heterocycles) were detected in biochars produced at >500 • C pyrolysis temperature (Ippolito et al., 2017). The negatively charged surface functional groups enable biochar to adsorb heavy metal cations through electrostatic attraction. Furthermore, the aromatic biochar surface abundant in π electrons could exert electrostatic attraction to electron-deficient metal cations through π-π * donor-acceptor interactions (Vithanage et al., 2017). Cations adsorbed through electrostatic attraction remain in the diffusion layer of biochar particles (i.e., outer-sphere adsorption) and are subject to leaching losses (Tan et al., 2015). Heavy metal ions in the soil solution can be adsorbed to biochar via ion exchange by replacing those cations (e.g., Ca 2+ , Mg 2+ , Na+, K + , and H + ) originally associated with the biochar surface functional groups (Fidel et al., 2018). This specific interaction (i.e., inner-sphere adsorption) is stronger than electrostatic attraction, yet the adsorbed metal cations are exchangeable, as influenced by the solution pH and ionic strength (Ding et al., 2014;Fidel et al., 2018). A lower pH and a higher solution ionic strength generally reduce the sorption of metal cations onto biochar through ion exchange, owing to the increased competition from other cations. Shim et al. (2015) reported that increasing the solution pH in the range of 3.0-6.0 significantly promoted the removal of Cu 2+ by Miscanthus straw-derived biochar from water. Metal cations may also form complexes with the surface functional groups and be strongly sorbed. Using positive matrix factorization analysis, Uchimiya et al. (2011a) validated surface complexation as the major mechanism for biochar sorbing Cd 2+ , Cu 2+ , Ni 2+ , and Pb 2+ via the ligand-like surface functional groups (e.g., carboxylic, hydroxyl, and phenolic groups). With the formation of biochar-metal complexes, Na + , K + , Ca 2+ , Mg 2+ , and other cations were released into soil, causing soil pH changes (Uchimiya et al., 2011a). Formation of organo-metal complexes on biochar surface and in pores was further evidenced by spectroscopic analyses (Kumar et al., 2018). Uchimiya et al. (2010a) reported that the addition of dissolved organic matter (DOM) substantially reduced the sorption of Cu 2+ , Cd 2+ , and Ni 2+ by biochars. Natural DOM contains surface ligands capable of complexing metal ions (Weng et al., 2002). The DOM released from biochar may reduce its sorption for heavy metal ions (Oustriere et al., 2016).
The abundance of surface functional groups as indicated by CEC is the most important trait that determines the capacity of biochar for stabilizing heavy metals through sorption-based interactions. Jiang et al. (2015) noticed that, among the biochars generated from rice straw through 200, 400, 500, and 600 • C slow pyrolysis, the 400 • C product had the highest CEC (60.65 cmol c kg −1 ) and demonstrated the greatest sorption capacity for aqueous Cu 2+ in batch sorption trials. Activation improves the SSA, aromaticity, and hydrophobicity of biochar but decreases its CEC, surface functional groups, and polarity (Uchimiya et al., 2010a;Shim et al., 2015;Yuan et al., 2019). Steam or CO 2 activation of willow-wood-derived biochar, for instance, increased the biochar's SSA from 11.4 to >510 m 2 g −1 but enhanced little its capability to reduce the ecotoxicity of a heavymetal-contaminated soil (Zn, Pb, Cr, Cu, Ni, and Cd) at 5 wt% amendment rate (Kołtowski et al., 2017).
Precipitation is equally important as sorption in biochar stabilizing soil heavy metals. In most cases, biochar amendment elevates soil pH by introducing additional alkalinity ( Table 2). Hydrolysis of heavy metals is promoted at higher pH, in which cationic metals react with hydroxyls in water to form metal hydroxide precipitates, reducing the concentration of watersoluble metal ions. Research has identified soil pH elevation as an important mechanism for biochar amendment to mitigate soil heavy metal contamination ( Table 2). Ippolito et al. (2017) noticed that addition of wood-based biochars (pH >9.1) at 5-15 wt% elevated the pH levels of mine land soils (pH <5.4) by 0.5-3.5 units and evidently reduced the 0.01 M CaCl 2 -extractable portions of the inherent Cu, Pb, Zn, and Cd. Other researchers observed similar soil pH elevation and soluble metal decreases when acidic soils were amended with pH 8.8-10.2 biochars (Park et al., 2011;Ahmad et al., 2012b;Houben et al., 2013;Zheng et al., 2015;Cui et al., 2016). In rice paddy soils, the elevated pH favored the formation of Fe/Al hydroxide precipitates; soluble metal ions could be encapsulated by co-precipitation in iron plaques deposited on rice roots (Zheng et al., 2012). Precipitation also occurs when cationic metal ions react with biochar-introduced carbonate, sulfate, and phosphate (Vithanage et al., 2017). Nearly all biochars contain carbonates (e.g., K 2 CO 3 , Na 2 CO 3 , CaCO 3 , and MgCO 3 ) that contribute to alkalinity and high pH (Cao et al., 2011;Jiang et al., 2012). Biochars derived from manures and sewage sludge usually contain substantial phosphates (Song and Guo, 2012). Mixing poultrylitter-, dairy-manure-, and sewage-sludge-derived biochars at 2-5 wt% with Pb-contaminated soils resulted in substantial decreases of soil 0.01 M CaCl 2 -extractable, physiologically based extraction test (PBET)-extractable, and toxicity characteristic leaching procedure (TCLP)-extractable Pb by forming pyromorphite [Pb 5 (PO 4 ) 3 Cl], hydroxypyromorphite [Pb 5 (PO 4 ) 3 (OH)], and Pb phosphate (Cao et al., 2011;Netherway et al., 2019). Sequential extraction analysis suggested that, in mining soils amended with wood-derived biochar that contained little P, precipitation of Cd 2+ was chiefly through carbonate formation, Pb 2+ through oxyhydroxide formation, and Cu 2+ and Zn 2+ through both carbonate and oxyhydroxide formation (Ippolito et al., 2017).
Clearly, biochar amendment facilitates the stabilization of heavy metals in contaminated soils through surface interactions (electrostatic attraction, ion exchange, and surface complexation) and (co-)precipitation (Figure 3). The carboxyl, hydroxyl, and phenolic groups on the biochar surface are particularly effective in binding cationic heavy metal contaminants (Uchimiya et al., 2011a).

Efficacy Variations
The efficacy of biochar amendment on immobilizing soil heavy metals is influenced by an array of factors, including the biochar source, amendment rate, metal species, soil type, and placement in soil (Ahmad et al., 2014;O'Connor et al., 2018). Biochar products manufactured from various biomass feedstocks through different pyrolysis conditions demonstrate great variations in pH, mineral ash content and composition, surface functionality, and CEC (Table 1) and consequently possess varied capability and effectiveness for stabilizing soil heavy metals ( Table 2). For example, Ding et al. (2014) reported that the maximum sorption capacity of sugarcane-bagasse-derived biochar for Pb 2+ in water (as soluble Pb 2+ in soil solution) decreased as the pyrolysis temperature of biochar production increased in the range of 250-600 • C. When mixing with contaminated rice paddy soil in field plots at 20 Mg ha −1 , rice-straw-derived biochar (pH 10.5; CEC 32.1 cmol c kg −1 ) reduced soil NH 4 NO 3 -extractable and rice-root-accumulated Cd, Pb, and Zn but increased As (arsenic) to significantly higher extents than soybean-strawderived biochar (pH 9.2; CEC 27.5 cmol c kg −1 ) prepared under the same carbonization conditions (Zheng et al., 2015); the concentration of As in rice shoot was more than tripled relative to the unamended control (Zheng et al., 2012). At 1, 5, and 10 wt% amendment rates, poultry-litter-based biochar reduced the Pb concentration of Indian mustard roots grown in pH 5.4 Australian soil from 3,243 to 1,276, 513, and 367 mg kg −1 , respectively, while greenwaste-based biochar reduced the level to 2,769, 2,299, and 1,196 mg kg −1 , respectively (Park et al., 2011), implicating the importance of biochar source and amendment dosage in developing biochar-facilitated soil remediation programs. A higher biochar amendment rate in the practical range (e.g., <5 wt%) typically yields greater results of heavy metal stabilization. Soil characteristics, in particular the clay and OC contents, and co-existing cations also affect the effectiveness of biochar amendment in stabilizing heavy metals. Uchimiya et al. (2011a) reported that addition of plant residues-derived 350 • C pyrolysis biochars at 10 wt% significantly enhanced the sorption of loamy sand (pH 5.6; OC 0.6%) for Cu 2+ and Pb 2+ but not Cd 2+ and Ni 2+ from a mixed metal solution (each 300 µM), whereas the sorption for Cu 2+ of clay soil (clay 50%, pH 9.2, OC 0.89%) from a mixed Cu 2+ /Cd 2+ /Ni 2+ solution (each 1.5 mM) was noticeably reduced with 10 wt% amendment of poultry-litter-derived 350 • C pyrolysis biochar. The researchers believed that biochar products from low-temperature-pyrolysis could not compete with soil clay particles for sorbing Cu 2+ . In the same tests, replacement with 700 • C-pyrolysis poultry litter biochar resulted in evident increases of Cu 2+ sorption by the clay soil (Uchimiya et al., 2010a). Biochar amendment affects plant growth. Lu et al. (2014) grew 60-day greenhouse amaranth (Amaranthus tricolor L.) in soil pots (pH 6.0; Cd 6.1 mg kg −1 ) amended separately with poultry-litter-and wood-based biochars both at 3 wt%. Relative to the control without biochar amendment, the poultry litter biochar amendment tripled the amaranth biomass yield, while the wood biochar amendment decreased such by 50%. Meanwhile, the total amount of amaranth-absorbed Cd reduced from 35.0 to 17.3 and 7.7 mg, respectively. In soils without chemical fertilization, manure-derived biochar furnishes additional nutrients and promotes plant growth, whereas plantresidue-derived biochar may immobilize available N and restrict plant growth (Guo, 2020). Furthermore, the effect of biochar amendment on stabilizing heavy metals may be initially strong but diminishes over time. In a 5-year field plot study, Cui et al. (2016) noticed that the leachability (by TCLP extraction), availability (by 0.01 M CaCl 2 extraction), and bioaccessibility (by simplified bioaccessibility extraction test) of Cu and Cd in contaminated paddy soil (pH 4.2) were significantly reduced by 3 wt% wood-based biochar amendment, 1 wt% apatite incorporation, and 0.2 wt% lime addition, yet the effect decreased gradually as time progressed.
In general, soil amendment with biochar at >2.0 wt% helps stabilize cationic heavy metals (e.g., Cd 2+ , Cu 2+ , Ni 2+ , Pb 2+ , and Zn 2+ ) and reduce the bioaccessible portion and the bioaccumulation of these toxic elements in soil. The efficacy is greater for biochar products possessing higher pH, CEC, and ash content levels, is more evident in acidic, coarse-textured, low OM soils, and becomes diminishing over time ( Table 2). Manurederived biochars are typically more efficient than wood-derived biochars in stabilizing soil heavy metals. Biochar amendment does not stabilize but helps mobilize anionic toxic elements [e.g., Cr in Cr 2 O 2− 7 and CrO 2− 4 , As in AsO − 3 and AsO 3− 3 , and Sb in Sb(OH) − 6 ] in soil. In biochar application, thorough and uniform incorporation of the amendment into contaminated soil is necessary to allow the biochar to get into direct contact with heavy metal contaminants. In all the reported research studies, biochar in <2 mm particles (mostly <1 mm) was fully mixed with the treated soils to achieve the remediation effect ( Table 2). The soil moisture content was controlled at >60% of the soil water holding capacity to secure efficient biocharheavy metal interactions that only occur in the soil solution phase. In combination with chemical fertilization, biochar amendment normally stimulates plant growth by improving the overall soil health (Guo, 2020). Practically, biochar amendment may be employed to temporarily rectify soils with slight to moderate heavy metal contamination. It may also be used jointly with phytoremediation to facilitate the "cleanup" of Cr (VI)-, As-, and Sb-contaminated soils. Long-term (e.g., >5 years) field research trials are warranted to validate the method feasibility. To date, merely one such study (Cui et al., 2016) was reported.
By reviewing the literature, Ahmad et al. (2014) concluded that biochars with more O-containing functional groups were desirable to immobilize inorganic and polar organic pollutants in soil and water. Accordingly, manure-derived biochar products from low pyrolysis temperature (e.g., <500 • C) may be more efficient in mitigating soil heavy metal contamination. Since biochar amendment does not remove heavy metals from soil, the method may only be applicable to slightly to moderately contaminated soils. In practice, appropriate biochar amendment rate and thorough soil incorporation have to be considered, with the remediation effect being regularly monitored.  Phenanthrene sorption on wood-based biochar was less evident; sorption on biochar was more evident in low-OC soils and reduced mineralization The source of biochar influenced its sorption performance Jiménez et al., 2018 Hard wood 450 and 600 • C pyrolysis pH 8.5-9.7 OC 76-78% SSA 4-39 m 2 /g Sandy clay loam (pH 6.2; OC 3.5%) and loamy sand (pH 4.8; OC 1.0%)

BIOCHAR FOR REMEDIATION OF ORGANIC-CONTAMINATED SOILS
Research has also demonstrated that biochar is capable of retaining and even promoting the decomposition of various organic contaminants in soils (Table 3). Biochar amendment may be a practical approach to mitigate soils polluted by pesticides, herbicides, antibiotics, PAHs, PCBs, petroleum hydrocarbons, and other POPs (Ahmad et al., 2014;Tan et al., 2015;Zama et al., 2018).

Mechanisms
After incorporation into contaminated soil with thorough mixing, biochar promptly interacts with organic contaminants and soil microorganisms. Organic contaminants are stabilized on the biochar surface and in pores and may be further decomposed by microbes as stimulated by biochar amendment. The porous, functional-group-abundant, and aromatic-C-condensed biochar surface is able to adsorb various organic compounds through different mechanisms. As organic pollutants are adsorbed by biochar, their concentrations in soil water are decreased, and bioaccessibility to soil organisms, including plant roots, is reduced. Meanwhile, biochar amendment enhances overall soil health by improving soil physical, chemical, and biological properties (Guo, 2020). In addition to furnishing mineral nutrients such as N, P, K, Ca, Mg, and S, biochar amendment introduces substantial amounts of biodegradable OC as soil microbial substrate. Following the initial perturbation period, biochar-amended soils generally demonstrate improved microbial community structure and promote microbial activity (indicated by soil respiration rate, soil enzyme activity, and soil microbial biomass) Irfan et al., 2019). Consequently, the microbial mineralization of many organic pollutants in soil is accelerated. Biochar is able to stabilize organic contaminants via a number of physical and chemical sorption mechanisms (Figure 4). Biochar interacts with organic molecules through van der Waals forces, the universal electrostatic attraction between nonpolar molecules (London dispersion forces) and between polar molecules (dipole-dipole forces) (Petrucci et al., 2007). The van der Waals forces, though fairly weak, increase with the molecules' surface area (e.g., higher molecular weight) and can be dominant as the particle size of biochar decreases (Yang and Evans, 2007). Biochar consists of C, O, H, N, and other mineral elements ( Table 1). It may interact with polar organic compounds by forming H bonds through the O-, N-, and H-containing functional groups (Figure 4). H-bonding is another form of electrostatic attraction stronger than van der Waals forces (Tan et al., 2015). The abundant functional groups in biochar become dissociated at pH > pH ZNC and protonated at pH < pH ZNC , engendering biochar with negative charges and positive charges, respectively (Fidel et al., 2018). The surface charges enable biochar to electrically attract polar organic molecules and ionized organic pollutants with counter charges (e.g., many pesticides, hormones, and antibiotics). This type of electrostatic attraction, as conventionally referred to, is generally stronger than H-bonding and van der Waals interactions. Intensified electrostatic attraction may lead to innersphere adsorption (specific surface interaction), in which ionized organic compounds react chemically with the biochar surface functional groups and are retained (Xu et al., 2011). In addition, the aromatic C in biochar possesses excess π-electrons, while many organic molecules, especially those containing O, N, S, P, Cl, and Br, are electron-deficit; biochar may interact with these organic contaminants through π-π * electron donor-acceptor interactions (Xie et al., 2014;Vithanage et al., 2017).
Biochar adsorbs non-ionic organic compounds through surface adsorption and/or partition. Smaller pores have higher surface energy, and therefore sorption of organic pollutants occurs first in biochar micropores (Uchimiya et al., 2010b). At low surface coverage [i.e., equilibrium solute concentration (Ce) to solute water solubility (Sw) ratio ≤0.2], non-linear, competitive adsorption of organic solutes into micropores in biochar's carbonized porous surface is dominant; at higher surface coverage (i.e., Ce/Sw >0.2), the adsorption shifts increasingly to linear, non-competitive partition into biochar's uncarbonized C moiety (Uchimiya et al., 2010b;Chiou et al., 2015;Zhao et al., 2019). Pore filling ("in-pore" surface adsorption) was suggested as an important mechanism for biochar adsorbing organic compounds (Figure 4). Research indicates that the extent of surface adsorption is proportionally related to biochar's SSA and surface porosity as well as aromaticity , while the capacity of partition adsorption is dependent on biochar's OC content and mineral ash content (Chiou et al., 2015;Zhao et al., 2019). The hydrophobic effect (avoidance of non-polar molecules and hydrophobic molecular moieties from contacting water) also facilitates the sorption of hydrophobic organic pollutants by biochar (Tan et al., 2015).
In addition, biochar contains more or less multivalent metal elements such as Fe, Al, Ca, and Mg on the surface (Bian et al., 2014). Polar and ionized organic compounds may form complexes with the metal ions and be deposited on the biochar surface or precipitated in soil. The sorption kinetics of many organic molecules on biochar follow the pseudo-second-order pattern, suggesting the importance of "chemisorption" (surface precipitation, Figure 4) in biochar-organic interactions (Tan et al., 2015). Xu et al. (2011) investigated the removal of methyl violet from water by two crop-residue-derived biochars as influenced by the solution pH and ionic strength and concluded that sorption of methyl violet by biochar was mainly through electrostatic attraction, specific interaction, and surface precipitation.
The surface properties of biochar, in particular SSA, pore size, pore volume, polarity, aromaticity, and hydrophobicity, have a predominant effect on biochar-organic compound interactions (Tan et al., 2015). In general, biochars produced at higher pyrolysis temperatures are greater in SSA, aromaticity, and hydrophobicity and lower in surface polarity due to the loss of O-and H-containing functional groups. Jiang et al. (2015) reported that, as the pyrolysis temperature was elevated from 200 to 400, 500, and 600 • C, the biochar products from rice straw increased in SSA from 7.2 to 19.7, 44.0, and 193.2 m 2 g −1 , respectively, and their sorption capacities for cyromazine in water also increased steadily, with the 600 • C biochar showing the highest organic sorption capacity. The mineral ash in biochar, especially the manure-derived products, may block the surface micropores and alter the SSA, hydrophobicity, polarity, and other surface properties. Zhang et al. (2013) observed that de-ashing by HCl-HF treatment increased the SSA value of pig-manure-based 350 • C (ash 45.3%) and 700 • C (ash 66.8%) pyrolysis biochars from 23.8 and 32.6 to 67.1 and 218.1 m 2 g −1 , respectively, but decreased the CEC level from 96.8 and 112.5 cmol c kg −1 to 5.6 and 8.7 cmol c kg −1 , respectively. As a result, the sorption capacities for the pesticides carbaryl and atrazine of pig-manure-based biochars were greatly enhanced, indicating the importance of biochar surface properties (Zhang et al., 2013). Clearly, different biochar products have varied sorption capacities and primary sorption mechanisms for stabilizing soil organic contaminants. Non-polar, hydrophobic organic contaminants such as PAHs and petroleum hydrocarbons are adsorbed by biochar mainly through pore filling, partition, and hydrophobic effect, while polar and ionized organic pollutants, including most pesticides and antibiotics, are adsorbed via Hbonding, electrostatic attraction, specific interaction, and surface precipitation (Figure 4).

Efficacy Variations
The efficiency of biochar amendment to stabilize organic pollutants and facilitate their elimination from soil is case specific, varying with biochar source, particle size, amendment dosage, pollutant chemical nature, and soil type ( Table 3). The biochars produced from canola straw, peanut straw char, soybean straw char, and rice hull through the same 350 • C pyrolysis processing, for example, demonstrated evidently different sorption capacities for removing methyl violet from water (Xu et al., 2011). Using 30-day incubation experiments, Koltowski et al. (2016) noticed that, at 5 wt% amendment rate, willow-, coconut shell-, and wheat straw-derived biochars reduced the freely dissolved (extractable by POM strips) and bioaccessible (extractable by silicon rods) fractions of PAHs in soils collected from three different industrial sites by 28-87% relative to the unamended controls. Steam activation greatly increased the biochars' SSA and pore volume and subsequently the sorption capacity for PAHs (Koltowski et al., 2016). Biochar itself may contain PAHs (up to 65 mg kg−1; Keiluweit et al., 2012;Quilliam et al., 2013), but mostly at insignificant, lowenvironmental-risk levels Weidemann et al., 2018). Amending sandy loam (pH 5.7; OC 0.54%) and clay (pH 5.8; OC 2.4%) soils with sugarcane-residue-derived biochar (pH 8.6; SSA 58.9 m 2 g −1 ; CEC 113.7 cmol c kg −1 ) at 5 wt% improved the sorption of 17α-ethinylestradiol (EE2, an estrogen hormone) on both soils and reduced their microbial mineralization (Wei et al., 2019). Furthermore, the sandy loam soil demonstrated a remarkably higher sorption efficiency for EE2 than the clay soil; the biochar amendment decreased the adsorption coefficient Koc of the former but increased that of the latter (Wei et al., 2019). When a strongly acidic (pH 3.2) mine spoil soil with mixed heavy metal and phenanthrene contamination was treated with 1 wt% biochar (N 2.1%; SSA 3.5 m 2 g −1 ) generated from stinging nettle (Urtica dioica L.), the phenanthrene degradation in soil microcosms was improved by 44% over a 56-day lab incubation period (Sneath et al., 2013). It was possible that the added biochar mainly functioned to stabilize the heavy metals and did not significantly adsorb phenanthrene. Nevertheless, the biochar amendment did ameliorate soil health and promoted microbial activity, as indicated by the stimulated sunflower (Helianthus annuus) grown in the amended soil pots (Sneath et al., 2013). Qin et al. (2016) reported that the enhanced microbial activity by biochar amendment was responsible for the 20% facilitation in petroleum hydrocarbon biodegradation in a petroleum spill site soil (clay loam; pH 6.5) amended with rice-straw-derived biochar (pH 8.9) at 2 wt% over a 180day lab incubation period. In addition to absorbing non-polar, hydrophobic organic compounds, biochar also showed high affinity for polar, hydrophilic organic pollutants. Jones et al. (2011) examined the effect of biochar amendment on leaching and biodegradation of the herbicide simazine in different soils. Two particle size fractions of commercial-wood-derived biochar were tested: coarse (2-10 mm) and fine (<2 mm). With addition at 0.5 and 5 wt% to soil, both biochar fractions suppressed the biodegradation of simazine and reduced its leaching in loamy sand (pH 4.8; OC 1.0%) and sandy clay loam (pH 6.2; OC 3.5%) soils, with more significant effects at a higher amendment rate, in finer particle size, and in greater-OC soils (Jones et al., 2011). The particle size of biochar influences the amendment efficacy through affecting the uniformity of biochar distribution in the treated soil. Denyes et al. (2013) noticed that soil incorporation of biochar by mechanical mixing was much more effective than hand or shovel mixing to achieve even distribution of biochar in soil and consequently enhance the retention of PCBs in 2.8 wt% biochar-amended brownfield soil and reduce PCB bioaccumulation in pumpkin (Cucurbita pepo) roots by >60%. Adsorption of pesticides, such as atrazine and simazine, by biochar may reduce the pesticidal efficacy of fieldapplied agrochemicals. To most surface-sprayed pesticides, the impact should be marginal.
In practice, wood-and plant-residue-derived biochars from higher pyrolysis temperatures and steam-activated biochars usually possess high SSA, porosity, and aromaticity levels and therefore may be selected with preference over manurederived biochars for sorption-based stabilization of organic contaminants. Over the long term, biochar amendment may facilitate the mineralization and the eventual elimination of organic contaminants in soil through improving soil microbial activities. The mechanisms through which biochar amendment improve soil health and soil microbial activities can be found elsewhere (Paz-Ferreiro et al., 2016;Guo, 2020).

CONCLUSIVE REMARKS
A notable number of studies have been conducted to investigate the feasibility of using biochar to remediate soils contaminated by heavy metals and various organic pollutants. The literature data indicate that biochar amendment is able to stabilize heavy metals and organic compounds in soil and mitigate soil contamination; the efficacy, however, is case specific, changing with biochar source, amendment dosage, soil type, pollutant species, and even biochar placement in soil. Differently sourced biochars vary significantly in physical and chemical quality properties and demonstrate diverse capacities and efficiencies for stabilizing soil heavy metals and organic contaminants. Biochar amendment does not remove heavy metals from soil; soil-incorporated biochar instead transforms the water-soluble and bioaccessible fractions of heavy metals into immobilized forms, reducing the bioavailability and the ecotoxicity of the toxic elements. This is achieved primarily through elevating the soil pH, introducing carbonates and phosphates, and enhancing surface sorption. Biochar stabilizes cationic heavy metals via electrostatic attraction, ion-exchange-based surface adsorption, surface complexation, and (co-)precipitation. For anionic forms of toxic elements, such as CrO 2− 4 , Cr 2 O 2− 7 , AsO − 3 , AsO 3− 3 , and Sb(OH) − 6 , biochar amendment generally improves their mobility in soil, suggesting that biochar may be used to facilitate phytoextraction-based remediation of Cr (VI)-, As-, and Sb-contaminated soils. The abundance of surface functional groups and the CEC of biochar determine its ability to adsorb heavy metals, while the pH/lime equivalence and the phosphate content control its capacity to precipitate heavy metal contaminants. This remark serves as a guideline for designing and selecting biochar products to mitigate soil heavy metal contamination. In general, manure-derived biochars at lower pyrolysis temperature were more efficient than plantresidue-derived biochars in stabilizing soil heavy metals. Soilincorporated biochar also stabilizes non-polar and polar organic pollutants through adsorption and reduces their ecotoxicity and bioaccumulation. Biochar amendment may promote the mineralization of organic contaminants in soil by enhancing soil health and microbial activity. Adsorption of biochar for non-polar organic contaminants is chiefly via pore filling, partition, and hydrophobic effect and for polar organic pollutants via hydrogen bonding, electrostatic attraction, specific surface interaction, and surface precipitation (chemisorption). The surface properties of biochar including SSA, microporosity, polarity, and aromaticity influence its capability for adsorbing organic pollutants. Plant-residue-derived biochars at higher pyrolysis temperature mostly outperformed manure-derived biochars in adsorbing organic contaminants. Overall, biochar amendment serves as a promising approach to stabilize soil heavy metals and organic contaminants and mitigate the hazardous soil pollution effects. Though the stabilization efficiency may not be comparable to other agents such as lime, phosphate salts, and activated carbon, biochar is more economically available and can furnish other environmental benefits such as carbon sequestration and soil health improvement when appropriately applied. Biochar amendment may be implemented in slightly polluted cropland to secure safe food production, in mine land restoration efforts to promote vegetation, and in soil bioremediation and phytoextraction projects to facilitate the process. The pollutant stabilization effect of biochar in field soils, however, may diminish over time. Field studies at larger scales are clearly needed to examine the long-term effect of biochar amendment on mitigating soil contamination under practical circumstances.
The application of biochar for improving soil health, facilitating soil remediation, and enhancing carbon sequestration remains currently at the fledgling stage. An emerging biochar industry with profitable commercial production and wide field application is foreseen. The research community has been focusing on validating the agricultural and the environmental benefits of soil-biochar amendment programs and developing biochar quality standards. Though Japan enacted the Soil Productivity Improvement Act in 1987 to sanction the practice of conditioning agricultural soils with charcoal and pot ash (Ogawa and Okimori, 2010), the US and the European countries have not established any governmental policies to regulate biochar use as a soil amendment. Production and application of biochar as a soil amendment may be subject to legislative regulations such as fertilizer policy, soil remediation and protection policy, urban planning policy, and biodiversity policy (van Laer et al., 2015). A sustainable biochar industry warrants systems thinking of feedstock supply security, bio-oil and producer gas standards, and national policies of renewable energy, CO 2 emission tax, and carbon sequestration credits (van Laer et al., 2015).

AUTHOR CONTRIBUTIONS
MG reviewed the literature, synthesized the information, and prepared the manuscript. WS helped with literature data collection and manuscript development and reviewed the manuscript. JT initiated the topic, conducted preliminary experiments and literature review, and examined the draft. All authors contributed to the article and approved the submitted version.

FUNDING
Financial support for compiling information was partially from the USDA-NIFA Evans-Allen project no. 1017883 and the USDA-NRCS Grant No. NR183A750010C001.