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Chemosphere 178 (2017) 466e478

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Chemosphere

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Review

Mechanisms of metal sorption by biochars: Biochar characteristics andmodifications

Hongbo Li a, Xiaoling Dong b, Evandro B. da Silva b, Letuzia M. de Oliveira b,Yanshan Chen a, *, Lena Q. Ma a, b, **

a State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Jiangsu 210023, Chinab Soil and Water Science Department, University of Florida, Gainesville, FL 32611, United States

h i g h l i g h t s

* Corresponding author. State Key Laboratory of Po** Corresponding author. State Key Laboratory of Po

E-mail addresses: chenyanshan@nju.edu.cn (Y. Ch

http://dx.doi.org/10.1016/j.chemosphere.2017.03.0720045-6535/Published by Elsevier Ltd.

g r a p h i c a l a b s t r a c t

� Biochar properties varied withincreasing pyrolysis temperature.

� Complexation and electrostaticinteraction are important mecha-nisms for As sorption.

� Complexation and reduction areimportant mechanisms for Cr and Hgsorption.

� Cation exchange and precipitationare important mechanisms for Cdand Pb sorption.

� Biochar have been modified toenhance its metal sorption capacity.

a r t i c l e i n f o

Article history:Received 10 January 2017Received in revised form11 March 2017Accepted 16 March 2017

Handling Editor: Patryk Oleszczuk

Keywords:BiocharHeavy metalContaminated waterSorptionFunctional groups

a b s t r a c t

Biochar produced by thermal decomposition of biomass under oxygen-limited conditions has receivedincreasing attention as a cost-effective sorbent to treat metal-contaminated waters. However, there is alack of information on the roles of different sorption mechanisms for different metals and recentdevelopment of biochar modification to enhance metal sorption capacity, which is critical for biocharfield application. This review summarizes the characteristics of biochar (e.g., surface area, porosity, pH,surface charge, functional groups, and mineral components) and main mechanisms governing sorptionof As, Cr, Cd, Pb, and Hg by biochar. Biochar properties vary considerably with feedstock material andpyrolysis temperature, with high temperature producing biochars with higher surface area, porosity, pH,and mineral contents, but less functional groups. Different mechanisms dominate sorption of As(complexation and electrostatic interactions), Cr (electrostatic interactions, reduction, and complexa-tion), Cd and Pb (complexation, cation exchange, and precipitation), and Hg (complexation and reduc-tion). Besides sorption mechanisms, recent advance in modifying biochar by loading with minerals,reductants, organic functional groups, and nanoparticles, and activation with alkali solution to enhancemetal sorption capacity is discussed. Future research needs for field application of biochar includecompetitive sorption mechanisms of co-existing metals, biochar reuse, and cost reduction of biocharproduction.

Published by Elsevier Ltd.

llution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, China.llution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, China.en), lqma@ufl.edu (L.Q. Ma).

H. Li et al. / Chemosphere 178 (2017) 466e478 467

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4672. Characteristics of biochar . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 468

2.1. Surface area and porosity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4682.2. pH and surface charge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4682.3. Functional groups . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4702.4. Mineral composition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 470

3. Mechanisms of metal sorption by biochar . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4703.1. Arsenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4713.2. Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4723.3. Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4733.4. Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4743.5. Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 474

4. Modification of biochar to enhance metal sorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4745. Future research directions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 476

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 476

1. Introduction

Heavy metals are ubiquitous in the environment, adverselyimpacting human health (J€arup, 2003). However, various anthro-pogenic activities including mining, smelting, fertilizer and pesti-cide application, and electronic manufacturing discharge haveincreased the amount of metal-containing wastewater into theaquatic environment, leading to water contamination with metals.To deal with metal-contaminated water, different methods havebeen suggested to remove metals from aqueous solution includingchemical precipitation, ion exchange, electrochemical treatment,and membrane technologies (Demirbas, 2008). Among themethods, biosorption technique is the most common and cost-effective. This is because biosorbents are environmentally friendlyand readily available in large quantities, and one of the most pop-ular biosorbents is biochar.

Biochar is a carbon-rich, fine-grained, and porous material. It isusually produced by thermal decomposition of biomass underoxygen-limited conditions at temperature <900 �C (Lehmann et al.,2006). It has received increasing attention due to its ability to storelarge amount of carbon, increase crop yield, reduce soil emission ofgreenhouse gases, improve soil quality, decrease nutrient leaching,and reduce irrigation and fertilizer requirements (Lehmann, 2007;Bird et al., 2008; Kimetu et al., 2008; Nguyen et al., 2009). Moreimportantly, due to the presence of highly-porous structure andvarious functional groups (e.g., carboxyl, hydroxyl, and phenolicgroups), biochar shows a great affinity for heavy metals (Mohanet al., 2007; Cao et al., 2009; Park et al., 2011). Much research hasexplored its ability for heavy metal removal from water (Ahmadet al., 2014; Mohan et al., 2014). Biochars are produced fromvarious feedstocks (wood bark, dairy manure, sugar beet tailing,pinewood, and rice husk) at different pyrolysis conditions (tem-perature, heating transfer rate, and residence time) to sorb metalsfrom water, including arsenic (As), cadmium (Cd), chromium (Cr),mercury (Hg), and lead (Pb) (Qian et al., 2015; Xie et al., 2015;Inyanga et al., 2016). For simplicity, metalloid As is grouped withmetals in this review.

Based on literatures, five mechanisms governing metal sorptionfrom water by biochar have been proposed (Ahmad et al., 2014;Mohan et al., 2014; Nartey and Zhao, 2014; Qian et al., 2015; Tanet al., 2015; Xie et al., 2015; Inyanga et al., 2016). They include:(1) electrostatic interactions between metals and biochar surface;(2) cation exchange between metals and protons or alkaline metals

on biochar surface; (3) metal complexation with functional groupsand p electron rich domain on the aromatic structure of biochar;(4) metal precipitation to form insoluble compounds; and (5)reduction of metal species and subsequent sorption of the reducedmetal species. The sorption mechanisms and capacity varyconsiderably with biochar properties and target metals. Recently,researchers reviewed biochar production technologies and metalremoval performance (thermodynamics, kinetics, isotherms, ca-pacity, and mechanisms) from water using biochar (Ahmad et al.,2014; Mohan et al., 2014; Nartey and Zhao, 2014; Qian et al.,2015; Tan et al., 2015; Xie et al., 2015; Inyanga et al., 2016). How-ever, most reviews provided sorption mechanisms for metals as agroup, lacking a comparison of the main mechanisms for removalof different metals. Since different metals show different species orvalence states at different solution pH conditions, the mainmechanisms for their sorption are different.

Compared to activated carbon, biochar is a promising adsorbentwith lower cost for metal removal from water. Metal sorption ca-pacities of biochar are 2.4e147, 19.2e33.4, 0.3e39.1,3.0e123 mg g�1 for Pb, Ni, Cd, and Cr, respectively (Inyanga et al.,2016). However, they are generally lower than that of activatedbiochar, which are 255 and 91.4 mg g�1 for Pb and Cd (Wilson et al.,2006). Therefore, biochars have been modified to enhance theirmetal sorption capacity by loading biochar with minerals, organicfunctional groups, reductants, and nanoparticles, and by activatingbiochar with alkali solution (Mohan et al., 2014). However, so far,there is no review on recent progress on biochar modificationexcept Mohan et al. (2014) who briefly discussed biochar modifi-cation by incorporating nanoparticles including magnetic particlesand carbon nanotubes.

In this review, we aimed to: 1) review the characteristics ofbiochar to better understand its efficiency in metal sorption, 2)discuss the dominant sorption mechanisms of individual metals bybiochar, 3) describe biochar modification to enhance its metalremoval from aqueous solutions, and 4) identify research needs andsuggest directions for future research. The novelty of the paper is tocompare the main mechanisms for removal of different metals bybiochar and to review recent progress on biochar modification toenhance metal sorption capacity. This review provides insight intobiochar materials and their capability to sorb different heavymetals, which is useful for future research and biochar fieldapplication.

H. Li et al. / Chemosphere 178 (2017) 466e478468

2. Characteristics of biochar

Physicochemical properties of biochar significantly influence itsability to sorb metals. Prior to exploring the mechanisms governingmetal removal by biochar, its properties need to be well charac-terized, including surface area, porosity, pH, surface charge, func-tional groups, and mineral contents.

2.1. Surface area and porosity

Surface area and porosity are the major physical properties thatinfluence metal sorption capacity of biochar. When biomass is py-rolyzed, micropores form in biochar due to water loss in dehydra-tion process (Bagreev et al., 2001). Biochar pore size is highlyvariable and encompasses nano- (<0.9 nm), micro- (<2 nm), andmacro-pores (>50 nm). Pore size is important for metal sorption,for instance, biochar with small pore size cannot trap large sorbate,regardless of their charges or polarity (Ahmedna et al., 2004).

Biochar porosity and surface area vary considerably with py-rolysis temperature. Studies show that elevated temperaturegenerally leads to larger pore size, thereby larger surface area(Fig. 1A, Table 1). With increasing temperature from 500 to 900 �C,porosity of biosolids biochar increased from 0.056 to 0.099 cm3 g�1,while surface area increased from 25.4 to 67.6 m2 g�1 (Chen et al.,2014). However, it should be noted that, in some cases, biochar

Pyrolysis temperature (

0 200 400 600

pH o

f bio

char

4

6

8

10

12

14 (B) pH

Pyrolysis temperature (

0 200 400 600

surfa

ce a

rea

(m2 g

-1)

0

20

40

60

80

200

400

600

800 (A) Surface area

Fig. 1. Surface area and pH of biochars produced from various feedstocks at p

produced at high temperature displays lower surface area andporosity. Chun et al. (2004) observed reduced surface area forbiochar produced from wheat straw at 700 �C compared to that at600 �C (363 vs. 438 m2 g�1). Similar result was reported by Jin et al.(2016) who compared biosolids biochar at 600 and 550 �C (5.99 vs.8.45 m2 g�1). At high temperature, biochar porous structure maybedestroyed or blocked by tar, leading to decreased surface area.

In addition to pyrolysis temperature, the composition of biocharfeedstock is also important. For example, the surface area ofmanure and biosolid biochar (5.4e94.2 m2 g�1) is much smallerthan that of plant biochar (112e642 m2 g�1) such as wheat, oakwood, corn stover, and pine needle (Table 1). Similarly, biosolidbiochar shows smaller porosity (0.053e0.068 cm3 g�1) than pine-needle biochar (0.076e1.90 cm3 g�1) produced at temperature of500e700 �C. Generally, biomass rich in lignin (e.g., bamboo andcoconut shell) develops macroporous-structured biochar, whilebiomass rich in cellulose (e.g., husks) yields a predominantlymicroporous-structured biochar (Joseph et al., 2007).

2.2. pH and surface charge

Similar to surface area and porosity, biochar pH varies withpyrolysis temperature and feedstock (Table 1). Generally, biochar isalkaline with some exceptions depending on feedstock (Table 1).Biochar produced from oak wood at 350 and 600 �C was acidic

oC)

800 1000conocarpus wasteswastewater sludge

oC)

800 1000

wheat straw

oak wood corn stovermunicipal biosolids maize straw

pine needlebroiler litter manure

poultry litter manureswine manure

yrolysis temperature ranging from 100 to 900 �C. Data are from Table 1.

Table 1Physico-chemical properties of biochars produced from various feedstocks under different pyrolysis temperatures.

Feedstock Temperature(�C)

Surface area(m2 g�1)

Porosity(cm3 g�1)

pH Atomic ratio Content of mineral elements (%) Reference

H/C O/C N/C K Ca Mg P

Wheat straw 300 116 0.55 0.29 Chun et al. (2004)400 189500 309 0.36 0.09600 438700 363 0.15 0.05

Oak wood 350 450 4.84 0.06 0.26 0.001 Nguyen et al. (2010)600 642 4.91 0.02 0.10 0.001

Corn stover 350 293 5.88 0.07 0.37 0.013 Nguyen et al. (2010)600 527 6.71 0.03 0.21 0.012

Municipal biosolids 400 5.49 8.46 1.01 0.120 Jin et al. (2016)450 7.21 9.74 0.87 0.120500 7.73 9.75 0.68 0.110550 8.45 10.5 0.58 0.110600 5.99 11.7 0.43 0.090

Maize straw 300 1.00 0.010 9.84 0.07 0.49 0.026 Wang et al. (2015g)450 4.00 0.010 10.5 0.06 0.35 0.023600 70.0 0.060 11.4 0.03 0.26 0.020

Pine needle 100 0.65 1.44 0.62 0.012 Chen et al. (2008)200 6.22 1.91 0.48 0.013250 9.52 1.08 0.40 0.012300 19.9 0.75 0.28 0.014400 112 0.044 0.45 0.17 0.013500 236 0.095 0.33 0.14 0.012600 207 0.076 0.26 0.10 0.010700 491 0.190 0.18 0.10 0.011

Broiler litter manure 350 59.5 0.000 Uchimiya et al. (2010)700 94.2 0.018

Municipal biosolids 500 25.4 0.056 8.81 0.48 0.45 0.075 0.85 5.93 1.47 1.82 Chen et al. (2014)600 20.3 0.053 9.54 0.22 0.30 0.064 0.85 6.27 1.55 1.88700 32.2 0.068 11.1 0.15 0.30 0.048 0.99 6.44 1.64 2.04800 48.5 0.090 12.2 0.03 0.17 0.026 0.93 6.58 1.66 1.93900 67.6 0.099 12.2 0.09 0.12 0.029 0.87 6.96 1.75 2.02

Poultry litter manure 400 5.4 0.003 9.50 3.88 2.83 1.73 1.22 Subedi et al. (2016)600 6.3 0.003 10.4 5.88 3.59 2.40 1.54

Swine manure 400 5.8 0.008 10.0 1.62 2.03 1.57 0.97 Subedi et al. (2016)600 10.6 0.011 10.4 3.53 2.89 2.13 1.55

Conocarpus wastes 200 7.37 0.06 0.41 0.011 0.04 4.34 0.34 0.08 Al-Wabel et al. (2013)400 9.67 0.04 0.18 0.012 0.05 5.18 0.40 0.09600 12.2 0.02 0.08 0.009 0.09 6.47 0.48 0.11800 12.4 0.01 0.06 0.011 0.12 6.75 0.78 0.13

Wastewater sludge 300 5.32 1.05 0.24 0.111 0.10 3.47 0.35 2.50 Hossain et al. (2011)400 4.87 0.76 0.17 0.102 0.11 4.17 0.43 2.80500 7.27 0.52 0.02 0.090 0.18 4.62 0.46 3.30700 12.0 0.30 0.00 0.050 0.20 5.35 0.54 3.60

Oak wood 200 4.60 0.014 0.13 0.39 0.04 0.03 Zhang et al. (2015a)400 6.90 0.021 0.38 1.18 0.15 0.06600 9.50 0.029 0.44 1.39 0.18 0.06

Wheat straw 200 6.11 1.42 0.58 0.018 Zhang et al. (2015b)400 10.8 0.63 0.20 0.012600 11.0 0.26 0.13 0.011

H. Li et al. / Chemosphere 178 (2017) 466e478 469

(4.84e4.91) (Nguyen et al., 2010). Similar low pH at 4.60 wasobserved for oak wood biochar produced at 200 �C, but at 400 and600 �C, the biochar was neutral to alkaline (6.90e9.50) (Zhanget al., 2015a). In addition, biochars from corn stover, wheat straw,and wastewater sludge at low temperature (200e400 �C) were alsoacidic with pH of 4.87e6.11 (Nguyen et al., 2010; Hossain et al.,2011; Zhang et al., 2015b).

Biochar pH increases with increasing pyrolysis temperature(Fig. 1B). Positive relationships have been observed between bio-char pH and pyrolysis temperature for biochars produced from oakwood (Nguyen et al., 2010; Zhang et al., 2015a), biosolids (Hossainet al., 2011; Chen et al., 2014; Jin et al., 2016), wheat, corn, andmaize residues (Nguyen et al., 2010;Wang et al., 2015g; Zhang et al.,2015b), manure (Subedi et al., 2016), and conocarpus wastes (Al-Wabel et al., 2013). Increasing temperature led to higher ashcomponent, which positively correlated with biochar pH (r ¼ 0.99;Jin et al., 2016), suggesting that ash component is a factor

contributing to biochar high pH. With increasing temperature from300 to 700 �C, contents of total base cations and carbonates inbiochar increased, contributing to increased pH (Yuan et al., 2011).In addition, disappearance of acidic functional groups such aseCOOH at high temperature is another contributor. With temper-ature increasing from 200 to 800 �C, basic functional groups onbiochar surface produced from conocarpus wastes increased from0.15 to 3.55 mmol g�1, while acidic functional groups decreasedfrom 4.17 to 0.22 mmol g�1, consistent with the increased biocharpH from 7.37 to 12.4 (Al-Wabel et al., 2013).

Another important property that influences metal sorption bybiochar is its surface charge. When biochar is applied to water formetal removal, solution pH strongly influences its surface charge.The point of zero charge (pHPZC) of biochar refers to the solution pHat which its surface net charge is zero. When solution pH is > pHPZC,biochar is negatively charged and binds to metal cations such asCd2þ, Pb2þ, and Hg2þ. When solution pH is < pHPZC, biochar is

H. Li et al. / Chemosphere 178 (2017) 466e478470

positively charged and binds metal anions such as HAsO42- and

HCrO4�. With increasing temperature from 500 to 900 �C, pHPZC of

biosolid biochar increased from 8.58 to 10.2 (Chen et al., 2014).Yuan et al. (2011) determined the zeta potential of biochars pro-duced from canola, corn, soybean, and peanut straw at 300, 500,and 700 �C. At solution pH of 3e7, all biochars were negativelycharged. However, compared to biochars produced at 300 and500 �C, those produced at 700 �C were less negatively charged,implying increased pHPZC with increasing pyrolysis temperature(Yuan et al., 2011). At higher temperature, the amounts of nega-tively charged functional groups on biochars (e.g., eCOOe, eCOH,and eOH) are reduced, resulting in less-negative surface chargesand increased pHPZC.

2.3. Functional groups

Functional groups such as carboxylic, amino, and hydroxylgroups play important roles in metal sorption. Pyrolysis tempera-ture and biochar feedstock are the two key factors controlling thequantities of functional groups on biochar surface. However, unlikegenerally increased surface area, porosity, pH, and pHPZC, theabundance of functional groups in biochar decreases withincreasing temperature, primarily due to higher degree of carbon-ization. With increasing temperature, atomic ratios of H/C, O/C, andN/C decrease (Table 1), suggesting decrease in abundances of hy-droxyl, carboxylic, and amino groups.

FTIR (Fourier Transform Infrared Spectroscopy) spectra havebeen widely employed to characterize the functional groups onbiochar surfaces. The FTIR spectra of functional groups in biocharsproduced at different temperatures are different. The FTIR spectraof wood and grass biochar changed when pyrolysis temperatureincreased from 100 to 700 �C (Keiluweit et al., 2010). Whencompared to the feedstock biomass, no significant difference inFTIR spectra was observed at low temperature of 100e200 �C,suggesting no change in functional groups. Dehydration of cellu-losic and ligneous components in biochar started at 300 �C(3500�3200 cm�1, wavenumber) whereas presence of lignin/cellulose-derived transformation products appeared at 400 �C(multiple peaks at 1600�700 cm�1). An increasing degree ofcondensation was observed at charring temperature �500 �C (in-tensity loss at 1650�1500 cm�1 relative to that at 885�752 cm�1)(Keiluweit et al., 2010).

During pyrolysis under increasing temperature, most functionalgroups of lignocellulosic materials are lost. Using FTIR photo-acoustic spectroscopy, Yuan et al. (2011) observed decreased in-tensity of peaks corresponding to carboxylic (eCOOH) and hydroxyl(eOH) groups as temperature increased from 300 to 700 �C forbiochars from canola, corn, soybean, and peanut straw. Comparedto those at 350 �C, similar loss of FTIR spectral features wasobserved for manure biochars produced at 700 �C (Cantrell et al.,2012). At higher temperature, these functional groups are ignited,leading to decreased amounts with increasing temperature.

In addition to FTIR, nuclear magnetic resonance (NMR) spectrahave been employed to characterize the functional groups of bio-char. Li et al. (2013) investigated the development of functionalgroups in rice straw and bran biochars produced at 100e800 �Cusing two-dimensional 13C NMR correlation spectroscopy. Biocharsfrom rice straw and bran went through dehydroxylation/dehydro-genation and aromatization process. Generally, with increasingtemperature, formation of O-alkylated groups and anomeric O-C-Ocarbons occured prior to the production of aromatic structures. Forbiochars produced at temperature <300 �C, aliphatic O-alkylatedcarbons were predominant, which were generally lost in biocharsproduced at >300 �C where aromatic structures were dominant (Liet al., 2013). Similarly, based on NMR spectroscopy, Zhang et al.

(2015b) observed decreased contribution of O-alkyl carbon from20-54% to 6.9e13% for wheat straw and lignosulfonate biochars astemperature increased from 200 to 600 �C, with alkyl carbon beingabsent in biochars produced at 600 �C.

2.4. Mineral composition

Mineral components including potassium (K), calcium (Ca),magnesium (Mg), and phosphorus (P) in biochar are also respon-sible for metal sorption from water. They can exchange or precip-itate with heavy metals and reduce their availability. Cao et al.(2009) proposed that precipitation of Pb phosphates such as Pbpyromorphite and hydropyromorphite was the main mechanismgoverning Pb sorption by dairy-manure biochar. During sorption ofmetal cations (Cd2þ, Cu2þ, Ni2þ, and Pb2þ) onto broiler litter bio-char, base cations such as Ca2þ, Mg2þ, Naþ and Kþ were releasedinto the solution from biochar via cation exchange (Uchimiya et al.,2010).

Both pyrolysis temperature and feedstock control the amountsof mineral components in biochar (Table 1). Total concentrations ofK, Ca, Mg, and P increase with increasing pyrolysis temperature forbiochar from biosolids (Hossain et al., 2011; Chen et al., 2014),manure (Subedi et al., 2016), conocarpus wastes (Al-Wabel et al.,2013), and oak wood (Zhang et al., 2015a). At higher temperature,biochar yield is lower, enriching minerals in biochar. However,water-soluble concentrations of mineral components behavedifferently from their total concentrations. Generally, during bio-char production, water-soluble concentrations of K, Ca, Mg, and Pincrease when heated at 200 �C but decrease beyond that tem-perature. This is probably due to increased crystallization as evi-denced by the formation of whitlockite [(Ca, Mg)3(PO4)2] orincorporation into the silicon structure at pyrolysis temperature of500 �C, which is less soluble (Shinogi, 2004; Cao et al., 2009). Be-sides temperature, feedstock is also an important factor influencingthe concentrations of mineral components in biochar. Generally, Pcontent in oak wood biochar (0.03e0.06%) is much lower thanbiosolid biochar (1.82e3.60%) (Hossain et al., 2011; Chen et al.,2014; Zhang et al., 2015a). Poultry litter and swine manure bio-chars generally contain higher K contents (1.6e5.9%) than biocharsfrom other materials (Subedi et al., 2016).

In summary, biochar properties vary considerably, mainlydepending on pyrolysis temperature and feedstock. However,temperature could have opposite effects on biochar properties,leading to opposite effects on metal sorption. For example, highpyrolysis temperature leads to higher surface area, providing moresites for metal sorption. However, it reduces the amounts of func-tional groups, which may lead to lower metal sorption viacomplexation between metals and functional groups. In addition,the impacts of biochar properties on metal sorption is metaldependent as different metals are sorbed via different mechanisms.To obtain biochar with desirable properties for metal removal,understanding the main mechanisms governing metal sorption isthe next step.

3. Mechanisms of metal sorption by biochar

Table 2 summarizes the sorption capacity and optimum solutionpH for metal sorption by biochar. The metal sorption capacity ofbiochar varies by 1e3 orders of magnitude, ranging from 1 to200 mg g�1 (Table 2). The pH for maximum metal sorption varieswith metals, as solution pH significantly influences both metalspeciation and surface charge of biochar. Change in solution pHimpacts the complexation behavior of functional groups such ascarboxyl, hydroxyl, and amino. For example, the ionization ofcarboxyl group is ~pH 3e4 (Pulido-Novicio et al., 2001). An increase

H. Li et al. / Chemosphere 178 (2017) 466e478 471

in pH makes carboxyl group deprotonated to effectively complexwith positively charged metals. For example, with solution pHincreasing from 3 to 7, straw biochars were more negativelycharged due to enhanced deprotonation of functional groups (Yuanet al., 2011).

Five mechanisms have been proposed to govern metal sorptionby biochar from aqueous solutions, i.e., complexation, cation ex-change, precipitation, electrostatic interactions, and chemicalreduction (Fig. 2). However, the role of each of mechanism plays foreach metal varies considerably depending on target metals. Previ-ous reviews described sorption mechanisms for metals as a groupwithout comparing the main mechanisms for sorption of differentmetals. Here, we reviewed the main mechanisms for sorption of As,Cr, Pb, Cd, and Hg.

Fig. 2. Conceptual illustration of heavy metal sorption mechanisms on biochar.

3.1. Arsenic

Arsenic is a carcinogenic metalloid, with concentrations innatural waters varying by several orders of magnitude dependingon source and local geochemical environment (Smedley andKinniburgh, 2002). Arsenic occurs in the environment in severaloxidation states (�3, 0, þ3, and þ5). In natural water, inorganicarsenate (AsV) and arsenite (AsIII) are predominant, with AsIIIbeing more toxic and mobile than AsV (Manning and Goldberg,1997). In aerobic environments, AsV is prevalent and exists asH2AsO4

� and HAsO42� at pH 3e11. Under reducing environment, AsIII

is dominant and exists as H3AsO30 at pH < 9.2 and H2AsO3

� atpH > 9.2 (Korte and Fernando, 1991; Lenoble et al., 2005).

Precipitation and reduction are minor mechanisms for Assorption by biochar. X-ray diffraction (XRD) analyses of pinewoodbiochar before and after AsV sorption showed no new peaks, sug-gesting no formation of new minerals and the precipitationmechanism was not important for AsV removal by biochar (Wanget al., 2015e). X-ray photoelectron spectroscopy analysis of a mag-netic biochar (produced from water hyacinth with chemical co-precipitation of Fe2þ/Fe3þ) after AsV sorption showed that ~89%

Table 2Sorption capacity and optimum pH of biochars produced from different feedstocks for m

Metal Feedstock (pyrolysis temperature)

AsIII empty fruit bunchrice huskpine wood (400e450 �C)pine bark (400e450 �C)oak wood (400e450 �C)oak bark (400e450 �C)

AsV empty fruit bunchrice husk

CrIII peanut straw (400 �C)soybean straw (400 �C)canola straw (400 �C)rice straw (400 �C)

CrVI sugar beet tailing (300 �C)coconut coir (produced 250, 350, 500, and 600 �C)

Cd daily manure (200 and 350 �C)oak bark (400e450 �C)

Pb sludge (550 �C)pine wood (400e450 �C)pine bark (400e450 �C)oak wood (400e450 �C)oak bark (400e450 �C)

Cu hardwood (450 �C)corn straw (600 �C)

Zn hardwood (450 �C)corn straw (600 �C)

Ni almond shell (650 �C)Hg Brazilian pepper (300, 450 and 600 �C)

soybean stalk (700 �C)

of the total As on biochar surface existed as AsV, suggesting littleAsV reduction (Zhang et al., 2016).

Unlike precipitation and reduction, complexation and electro-static interactions are important mechanisms for As sorption bybiochar. Samsuri et al. (2013) compared AsV and AsIII sorptionbetween two biochars from empty fruit bunch and rice husk.Although the surface area of empty fruit bunch biochar wassignificantly lower (1.9 m2 g�1) than that of rice husk biochar(25.2 m2 g�1), its As sorption capacity (5.5 mg g�1 AsV and18.9 mg g�1 AsIII) was similar to rice husk biochar (7.1 mg g�1 AsVand 19.3 mg g�1 AsIII) (Table 2). This could be explained by higherH/C and O/C molar ratios (0.08 and 0.61) for empty fruit bunch

etal sorption from aqueous solutions.

pH Sorption capacity (mg g�1) Reference

8.0 18.9 Samsuri et al. (2013)8.0 19.33.5 1.20 Mohan et al. (2007)3.5 12.23.5 5.853.5 7.406.0 5.50 Samsuri et al. (2013)6.0 7.104.0 25.0 Pan et al. (2013)4.0 17.24.0 14.64.0 14.02.0 123 Dong et al. (2011)3.0 31.1, 10.9, 7.90, and 4.10 Shen et al. (2012)e 31.9 and 51.4 Xu et al. (2013)5.0 5.40 Mohan et al. (2007)5.0 30.9 Lu et al. (2012)5.0 4.13 Mohan et al. (2007)5.0 3.005.0 2.625.0 13.15.0 6.79 Chen et al. (2011)5.0 12.55.0 4.54 Chen et al. (2011)5.0 11.06.0 20.0 Kılıç et al. (2013)6.0 24.2, 18.8, and 15.1 Dong et al. (2013)7.0 0.67 Kong et al. (2011)

Fig. 3. Mechanisms of CrIII (A) and CrVI (B) sorption by biochar.

H. Li et al. / Chemosphere 178 (2017) 466e478472

biochar than rice husk biochar (0.05 and 0.37), suggesting higheramounts of oxygenated functional groups such as carboxyl andhydroxyl groups, which could counterbalance the low As sorptiondue to low surface area (Samsuri et al., 2013). This suggests that Ascomplexation with functional groups controls As sorption by bio-char, which is supported by the shift in FTIR spectra peaks offunctional groups including hydroxyl, carboxyl, and C-O ester ofalcohols (Samsuri et al., 2013). Following As loading, absorptionbands corresponding to hydroxyl group (3386 cm�1), CeH groups(2925 cm�1), COOe groups (1576 cm�1), CH2-groups (1369 cm�1),and CeO ester of alcohols, carboxylic acid groups and carboxylicacids (1020e1300 cm�1) shifted, while new peaks appeared whichwere characteristics of adsorbed As compounds. Zhang et al.(2015e) used biosolid biochar produced at different temperatures(300e600 �C) to sorb AsIII, showing that biochar at 600 �C had alower sorption capacity (0.95 mg g�1) than that at 300 �C(2.84 mg g�1) due to loss of oxygenated functional groups. Wanget al. (2015d) measured As sorption by 12 biochars from 4 feed-stocks at 3 different temperatures (300, 450, and 600 �C), showingsimilar results that sorbed AsV generally decreased with increasingpyrolysis temperature. These data confirm the importance offunctional groups for As sorption by biochar.

In addition to complexation, electrostatic interactions is anotherimportant mechanism for AsV sorption by biochar. Wang et al.(2015e) used pinewood biochar produced at 600 �C (pHPZC > 7)to sorb AsV from water at pH 7, which showed a maximum AsVsorption of 0.3 mg g�1. At solution pH 7, AsV mainly existed asHAsO4

2� while the biochar surface was positively charged, withsome functional groups being protonated as solution pH was <pHPZC. AsV oxyanions interact with the positively charged func-tional groups by electrostatic attraction. At lower solution pH,biochars are more positively charged with higher degree of pro-tonation of functional groups than those at high solution pH,thereby having greater ability to attract AsV oxyanions via elec-trostatic interactions. Wang et al. (2016b) reported that AsV sorp-tion increasedwith decreasing solution pH from9 to 2 by pinewoodbiochars with pHPZC > 10. Similar pH effects on AsV sorption wereobserved for Ni/Mn oxide-modified pinewood biochars (Wanget al., 2016a).

In short, complexation and electrostatic interactions areimportant mechanisms for As removal by biochar, with functionalgroups being the dominant property governing As sorption. Pro-ducing biochar at lower pyrolysis temperature using suitablefeedstock can help to obtain desirable biochar having higheramounts of functional groups to remove As from water.

3.2. Chromium

Chromium (Cr) exists in many valence states ranging from �2to þ6, with CrIII and CrVI being the major two oxidation states innatural waters (Rai et al., 1989). CrVI is highly soluble and mobile inaqueous solution and is of significant environment concern becauseof its carcinogenic, mutagenic, and teratogenic behavior in biolog-ical systems (Fendorf et al., 2000). CrIII is usually considered as anessential micronutrient for humans, being ~300 times less toxicthan CrVI. Although CrIII is considerably less toxic than CrVI, itsdisposal as a soluble species in natural waters may pose serioushealth risks because it can be oxidized to CrVI in the environment(Fendorf et al., 2000). Under oxidizing conditions, the principal CrVIspecies are HCrO4

�, CrO42- and Cr2O7

2-. Overall, Cr2O72- and HCrO4

�

dominate at pH < 6.0 while CrO42- dominates at pH > 6.0 (Richard

and Bourg, 1991). In low Eh environments, the main CrIII speciesare Cr3þ, Cr(OH)2þ, Cr(OH)3(s) and Cr(OH)4- . At pH < 3.6, Cr3þ is theprevalent species.

Studies of Cr removal using biochar are mainly focused on CrVI,

with few on CrIII (Wnetrzak et al., 2013; Yang et al., 2013; Pan et al.,2013). However, based on the limited literatures, threemechanismsare responsible for CrIII sorption by biochar: (1) complexation withoxygen-containing functional groups, (2) cation exchange, and (3)electrostatic attraction between positively charged CrIII ions andnegatively charged biochar (Fig. 3A).

Biochars prepared from crop straws were investigated for CrIIIsorption, with sorption capacity following the order ofpeanut > soybean> canola > rice (0.48, 0.33, 0.28, 0.27mmol kg�1),being consistent with the amounts of oxygen-containing functionalgroups (1.34, 1.13, 0.80, 0.63 mmol g�1). The data suggest that CrIIIcomplexation with functional groups is important for its sorptionby biochar (Pan et al., 2013). This was evidenced by peak shifts offunctional groups in FTIR spectra for biochar following CrIII sorp-tion, consistent with aromatic C]C ring stretching, phenolic OHregion, and aliphatic CeH stretching (Pan et al., 2013). By separatingrice straw biochar into biochar colloids (<2 mm) and residues(>2 mm), Qian et al. (2016) observed significantly higher CrIIIsorption capacity for biochar colloids than residues, consistent withhigher amounts of oxygen functional groups in biochar colloids,implying the role of complexation in CrIII sorption.

However, comparing CrIII sorption by pig manure biochar pro-duced at different pyrolysis temperatures, Wnetrzak et al. (2013)observed significantly higher sorption capacity of biochar pro-duced at 600 �C than 400 �C at solution pH of 4e5 (21e26 vs.17e19 mg g�1). Similarly, Qian et al. (2016) observed significant

H. Li et al. / Chemosphere 178 (2017) 466e478 473

increase in CrIII sorption by biochar colloids with increasing tem-perature from 100 to 400 �C. FTIR spectra peak position and in-tensity showed no significant changes following CrIII sorption bybiosolid biochar, suggesting little complexation of CrIII with bio-char functional groups (Chen et al., 2015b). They observed releaseof cations (Ca2þ and Mg2þ) into solution from biochar during CrIIIsorption, and the released cations correlated well with sorbed CrIII,suggesting cation exchange is an important sorptionmechanism. Inaddition, during CrIII sorption, slight increase in solution pH wasobserved (Chen et al., 2015b; Qian et al., 2016), possibly due torelease of CaO and MgO from biochar into solution as supported byincreased Ca and Mg concentration in solution following CrIIIsorption.

Sorption of CrIII is pH dependent, with increasing sorption withincreasing solution pH at 2.5e5.0 (Pan et al., 2013; Yang et al., 2013).At lower solution pH, higher concentration of Hþ inhibits cationexchange between CrIII and minerals in biochar. However, sorptionof CrIII through electrostatic attraction cannot be ruled out. At so-lution pH of 2.5e5.0, biochars from crop straw were negativelycharged based on their negative zeta potential values, while CrIIIwas mainly positively charged (Cr3þ, CrOH2þ, and Cr(OH)2þ) (Panet al., 2013). As solution pH increases, biochars become morenegatively charged, showing higher electrostatic attraction abilityfor CrIII.

Different from CrIII, two mechanisms have been proposed forCrVI sorption: (1) electrostatic attraction between negativelycharged CrVI species and positively charged biochar; and (2)reduction of CrVI to CrIII mainly by oxygen-containing functionalgroups such as carboxyl and hydroxyl groups and subsequent CrIIIcomplexation with functional groups on biochar (Fig. 3B). Amongthe mechanisms, CrVI reduction to CrIII followed by CrIIIcomplexation is a major sorption mechanisms for CrVI. Dong et al.(2011) observed that sugar beet tailing biochar was effective in CrVIsorption, which was mainly through CrVI reduction to CrIII andthen complexation by hydroxyl and carboxyl groups on biochar,with maximum sorption of 123 mg g�1 at pH 2.0 (Table 2). Usingcoconut coir biochar produced at 250, 350, 500, and 600 �C, Shenet al. (2012) observed that at pH 2.0, CrVI sorption decreasedsharply from 31.1 to 4.10 mg g�1 with increasing temperature from250 to 600 �C, consistent with sharp decrease in acidic functionalgroups (carboxyl, lactonic, and phenolic) from 1.78 to0.12 mmol g�1. Similar results that low-temperature biochars havehigher CrVI sorption capacity have been reported by others (Hanet al., 2016; Zhang et al., 2013; Zhou et al., 2016). Using waste-water sludge biochar produced at 300e600 �C, Zhang et al. (2013)observed significantly higher CrVI sorption capacity for biochar at300 �C (208 mg g�1) compared to that at 400e600 �C(36.6e141 mg g�1) at pH 2. In addition, longer pyrolysis residencetime decreases CrVI sorption capacity. With pyrolysis residencetime increasing from 1 to 2 h, municipal wastewater sludge bio-chars produced at 400e600 �C showed decreased CrVI sorptionfrom 69.0e118mg g�1 to 19.6e45.2 mg g�1 due to loss of functionalgroups (Zhang et al., 2013). The data confirm that CrVI reductionfollowed by CrIII complexationwith functional groups on biochar isthe main mechanism for CrVI sorption. Reduction of CrVI to CrIIIhas been confirmed based on spectra analysis of biochar surface byX-ray photoelectron spectroscopy where CrIII and CrVI co-existed,with CrIII being the dominant (93%) species of sorbed Cr (Donget al., 2011; Zhang et al., 2013).

Mohan et al. (2011) further studied CrVI sorption ability of oakwood and oak bark biochars, with wood containing higher amountof lignin, celluloses and hemicelluloses. When oak wood was sub-jected to pyrolysis, the biochar contained substantial amount ofoxygen (8e12%). Oxygen-containing compounds includingbyproducts catechol, substituted catechol, unsaturated

anhydrosugars, and diols generally are effective in reducing CrVI toCrIII (Mohan et al., 2011; Ahmad et al., 2014). Furthermore, thereduction mechanisms can be separated into direct and indirectpathways. The direct reduction occurs in the aqueous phase bysolubilized biochar components in aqueous solution (Dong et al.,2014). The indirect reduction occurs on solid biochar surfacewhere CrVI is bound to biochar surface due to electrostatic attrac-tion prior to its reduction to CrIII (Zhou et al., 2016). In short, unlikeother metals, biochar enhances CrVI sorption via reduction of toxicCrVI to less toxic CrIII (Fig. 3B). Using biochar produced at lowtemperature with high amounts of functional groups can facilitateCrVI removal.

3.3. Cadmium

Divalent metal cations such as Cd and Pb have a strong tendencyto be hydrated in aqueous solution, which is pH dependent. Due totheir similarity in aqueous solution, divalent metals share similarsorption mechanisms, i.e., cation exchange, surface complexation,precipitation, and electrostatic interactions. Among divalentmetals, sorption of Cd and Pb by biochar has been studied the most.

Cadmium has little tendency to be hydrolyzed at pH < 8.0, but atpH > 11.0, all Cd exists as hydrxo-complex. In natural aquatic so-lution, Cd2þ is the predominant species. Mechanisms for Cd sorp-tion by honey mesquite, cordgrass, and loblolly pine biochars fromaqueous solutionwere reported by Harvey et al. (2011). These plantbiochars were grouped into two groups based on their cation ex-change capacity: low and high. For the high capacity group, cationexchange was the predominant mechanism for Cd sorption. Flowcalorimetry experiment was used to study the behavior of K and Cdto replace Na-saturated biochar. The shape and duration of heatsignals of Cd exchange for Na was similar to that of K exchange forNa based on flow calorimetry, indicating that cation exchange is thedominant mechanism (Harvey et al., 2011). Zhang et al. (2015d)showed that released cations (sum of K, Ca, Na, and Mg) fromwater hyacinth biochar was almost equal to the amount of sorbedCd, suggesting the dominant role of cation exchange in Cd sorptionby biochar. Trakal et al. (2014) tested Cd sorption ability of biocharsfrom waste agro-materials (nut shells, plum stones, wheat straws,grape stalks, and grape husks), showing that grape stalk biocharwas more effective in Cd sorption than plum stone biochar (0.45 vs.0.04 mmol kg�1), consistent with its higher cation exchange ca-pacity than plum stone biochar (402 vs. 121 mmol kg�1), againsuggesting cation exchange as the predominant mechanism.

Comparison of FTIR spectra of biochar before and after Cdsorption showed insignificant shift in peaks of carboxyl groupsfollowing Cd sorption (Trakal et al., 2014; Chen et al., 2015a), sug-gesting that Cd complexation with carboxyl groups is a minormechanism in Cd sorption. However, for honey mesquite, cord-grass, and loblolly pine biochars with low cation exchange capacity,calorimetric heat signals were almost three times higher than thatfor high capacity group (Harvey et al., 2011), which was inconsis-tent with cation exchange mechanism. Although cation exchangecannot be ruled out, there is strong evidence that complexationwith oxygenated functional groups and electron-rich domains ongraphene-like structures is an important mechanism for Cd sorp-tion by biochar with low cation exchange capacity.

Due to relatively high soluble concentrations of carbonate andphosphate in dairy manure biochar, precipitation was suggested asthe mainmechanism for Cd sorption (Xu et al., 2013). With increasein temperature from 200 to 350 �C, sorption capacity of Cdincreased from 31.9 to 51.4 mg g�1 mainly due to increase inminerals especially soluble CO3

2- in biochar (2.52 vs. 2.94%)(Table 2). Visual MINTEQ modeling coupled with FTIR experimentshowed that 88% of Cd sorbed onto dairy manure biochar produced

H. Li et al. / Chemosphere 178 (2017) 466e478474

at 350 �C was attributed to precipitation of metal-phosphate and-carbonate, with 12% of Cd sorption being from Cd-p bonding (Xuet al., 2013). While <3% of the precipitation resulted from metalphosphate due to low soluble P in biochar produced at 350 �C,carbonate precipitates accounted for up to 98% of Cd beingprecipitated. For biochar produced at 200 �C, precipitationaccounted for 100% sorption of Cd, with 22% from phosphate pre-cipitates due to relatively high content of soluble P with carbonateprecipitates accounting for 78% (Xu et al., 2013). X-ray diffractionanalyses confirmed the formation of Cd carbonate and phosphateminerals in water hyacinth biochar following Cd sorption (Zhanget al., 2015d). In addition, based on FTIR spectra, Trakal et al.(2014) showed that the peaks of CO3

2- in high ash content biocharfrom grape stalks and husks shifted following Cd sorption probablydue to surface precipitation of Cd carbonates.

3.4. Lead

In natural aquatic system, Pb is present mainly as Pb2þ andPb(OH)þ at pH < 5.5, Pb(OH)2� at pH 5.5e12.5 and Pb(OH)42- atpH > 12.5 (Ucun et al., 2003). Various biochars including biosolids,dairymanure, oakwood, oak bark, and bagasse have been tested fortheir ability to sorb Pb from aqueous solutions (Cao et al., 2009;Mohan et al., 2007; Ding et al., 2014). Mechanisms governing Pbsorption by biosolids biochar were reported by Lu et al. (2012),including cation exchange, complexation, and precipitation.

Among the three mechanisms, cation exchange with Ca and Mgis the main one contributing to Pb sorption by biosolid biochar,which explained 40e52% of the Pb sorption at pH 2e5, while ex-change with K and Na contributed 4.8e8.5%, with Pb complexationwith carboxyl and hydroxyl groups contributing 38e42% (Lu et al.,2012). Cation exchange is also the main mechanism for Pb sorptionby oak wood and oak bark biochars, which is supported by the factthat the amount of Pb sorbed onto biochar was similar to that of thecations released (Mohan et al., 2007). Ding et al. (2014) also re-ported that cation exchange accounted for 62% of Pb sorbed bybagasse biochar produced at 500 �C.

However, for dairy manure biochar produced at 200, 250 and350 �C, due to the high content of phosphate and carbonate, pre-cipitation as Pb phosphate and Pb carbonate minerals (84e87%)was the predominant mechanism, which was confirmed bychemical speciation, X-ray diffraction, and FTIR spectroscopy data,while complexation with functional groups may contribute to therest (Cao et al., 2009). By separating biochar (dairy manure and ricestraw) into organic and inorganic fractions, Xu et al. (2014) showedthat the Pb sorption capacity of organic fraction wasonly ~ 1 mg g�1, while it was >300 mg g�1 for inorganic fraction,suggesting limited contribution of Pb complexation with organicfunctional groups to Pb removal, while cation exchange and Pbprecipitation were the dominant mechanisms. For manure biochar,precipitation with phosphate contributed more to Pb sorption thanthat with carbonate (68 vs. 32%), while the opposite was true forstraw biochar (36 vs. 64%).

In short, cation exchange, complexation, and precipitation arethe three main mechanisms responsible for Cd and Pb sorption bybiochars. However, their sorption mechanisms depend on thecharacteristics of biochar, which is affected by feedstock, pyrolysistemperature, and solution pH.

3.5. Mercury

Mercury exists in three different oxidation states: 0, þ1, and þ2,with divalent Hg as the most common species in the environment(Loux, 1998). In aqueous solution, HgII exists as Hg2þ at pH < 3.0,and HgOHþ and Hg(OH)2 at pH 3.0e7.0 (Das et al., 2007).

Complexation with carboxylic and phenolic hydroxyl groups orgraphite-like domainwas the dominantmechanism for Hg sorptionby Brazilian pepper biochar (Dong et al., 2013). With increasingpyrolysis temperature from 300 to 600 �C, Brazilian pepper biocharshowed decreasing Hg sorption capacity from 24.2 to 15.1mg g�1 atpH 7.0 (Table 2). For biochar produced at 300 and 450 �C, 23e31%and 77�69% of sorbed Hg was associated with carboxylic andphenolic hydroxyl groups, while for biochar at 600 �C, 91% of sor-bed Hg was associated with a graphite-like domain on an aromaticstructure, with the rest being associated with phenolic hydroxylgroups. The decreased Hg sorption capacity by biochar is probablydue to reduced carboxylic and phenolic hydroxyl groups withincreasing temperature. Xu et al. (2016) compared Hg sorption bytwo biochars from bagasse and hickory chips, showing that surfacecomplexation was the most important mechanism. However,different biochar sorbs Hg via different complexation mechanisms.As evidenced by XPS spectra, Hg sorption by bagasse biochar wasmainly attributed to the formation of (eCOO)2Hg and (eO)2Hg.Following blocking eCOOH and eOH functional groups usinganhydrous methanol and methanol, sorption capacity of Hg bybagasse biochar decreased 18% and 38%. However, the blocking hadlittle effects on Hg sorption by hickory chip biochar, since Hgsorptionwasmainly resulted from the p electrons of C]C and C]Oinduced Hg-p binding (Xu et al., 2016). In addition to hydroxyl andcarboxylic groups, Hg complexation with thiol groups was alsoobserved based on Hg extended X-ray absorption fine structureanalyses (Liu et al., 2016). Thirty-six biochars from different feed-stocks were tested for their ability to sorb Hg, with high S biocharshowing binding of Hg with S, while Hg was mainly bound to O andCl in biochars with low S content (Liu et al., 2016).

Besides complexation, chemical reduction is responsible for Hgsorption. Kong et al. (2011) investigated Hg sorption by soybeanstalk biochar. Besides cation exchange, complexation, and Hg(OH)2precipitation, they also proposed that Hg2þ was reduced to Hg2Cl2in presence of Cl�, which was then precipitated on biochar surface.However, there was no direct evidence to prove the presence ofHg2Cl2 on biochar. Reduction of Hg2þ to Hgþ by phenol groups or pelectrons was observed during the removal of Hg2þ by biocharbased on XPS analyses (Xu et al., 2016). Though numerous studieshave investigated Hg sorption by activated C, there is no consistentevidence regarding Hg precipitation. Many researchers haveattributed Hg sorption on activated C to the presence of Hg(OH)2due to its existence as an uncharged soluble hydroxide salts, whileothers have insisted that if Hg precipitates with Cl� during sorption,then it is likely to be Hg2Cl2 (Lloyd-Jones et al., 2004). Hence, morestudies are needed to prove whether Hg2þ reduction to Hgþ is animportant mechanism to control Hg sorption by biochar.

4. Modification of biochar to enhance metal sorption

Though biochar has ability to sorb metals from water, its ca-pacity is usually lower compared to other common biosorbentssuch as activated C. Therefore, recent studies havemodified biocharto enhance its metal sorption capacity. For example, efforts havebeen made to increase its surface area, porosity, pHPZC, and/orfunctional groups. Approaches to modify biochars include loadingwith minerals, reductants, organic functional groups, and nano-particles and activation with alkali solution.

Biochar modification includes loading biochar with differentminerals such as hematite (g-Fe2O3), magnetite, zero valent Fe,hydrous Mn oxide, calcium oxide, and birnessite (Table 3). Theloading can be achieved before, during, or after pyrolysis of thefeedstock. Wang et al. (2015e) synthesized a magnetic biochar bypyrolyzing a mixture of hematite mineral and pinewood, therebyincorporating g-Fe2O3 onto biochar surface and serving as

Table 3Modification of biochars with various materials to enhance its absorption ability for heavy metal removal from aqueous solutions.

Type modification Metal Sorption capacity/removalefficiency

Mechanisms Reference

Minerals magnetic biochar ofhematite þ pinewood

AsV from 265 to 429 mg kg�1 g-Fe2O3 particles as sorption sitesvia electrostatic interactions

Wang et al. (2015e)

magnetic biochar fromiron chloride þ peanut hull

CrVI 1-2 orders higher g-Fe2O3 particles as sorption sitesvia electrostatic interactions

Han et al. (2016)

iron þ empty fruit bunch AsV, AsIII from 5.5 and 18.9 to 15.2and 31.4 mg g�1

As complexation with FeOH2þ

and FeOH2þ groups

Samsuri et al. (2013)

iron þ rick husk AsV, AsIII from 7.1 and 19.3 to 16.9and 30.7 mg g�1

As complexation with FeOH2þ

and FeOH2þ groups

Ca/Fe þ rice husk AsV AsV removal efficiency increasedfrom 25% to 58e95%

metal precipitation andelectrostatic interactions

Agrafioti et al. (2014)

amorphous hydrous Mnoxide þ pine wood

Pb from 6.4 to 98.9% at pH 5.00 increased surface hydroxylsand decreased pHPZC

Wang et al. (2015b)

MnCl2$4H2O or birnessiteþ þ pine wood

Pb, AsV from 0.2 to 0.59e0.91 for Pband 2.35 to 4.91e47.1 for AsV

strong AsV and Pb affinity forbirnessite particles

Wang et al. (2015c)

KMnO4 þ hickory wood Pb, Cd by 2.1 and 5.9 times to153 and 28.1 mg g�1

more surface oxygen-containingfunctional groups andlarger surface area.

Wang et al. (2015a)

Zn þ pine cones AsIII increased from 66.1 to 87.6% increased surface hydroxylsand decreased pHPZC

Van Vinh et al. (2015)

Reductant Zero valent iron þ bamboo Pb, CrVI, AsV enhanced Pb, CrVI, and AsVremoval efficiency from23.9, 0.0, and 1.0% to 90e100,25e40, 20e95%

metal reduction andsurface sorption

Zhou et al. (2014)

Na2SO3/FeSO4 þpeanut straw CrVI efficient CrVI removal enhanced CrVI reductionand CrIII complexation

Pan et al. (2014)

Organic functionalgroups

polyethylenimine þ rice husk CrVI from 23.1 to 436 mg g�1 increased amino groups Ma et al. (2014)chitosan þ magneticbiochar þ Eichhornia crassipes

CrVI from 30 to 120 mg g�1 Increased functional groups Zhang et al. (2015c)

b-cyclodextrinechitosan þwalnut shell

CrVI from 27 to 93% enhanced surface area,porosity and thermal stability

Huang et al. (2016)

Decorated withnano-particles

Graphene þ wheat straw Hg from 71 to 80% larger surface area, more functionalgroups, greater thermal stability

Tang et al. (2015)

Zn þ sugarcane bagasse CrVI by 1.2e2.0 times increased surface area and pore volume Gan et al. (2015)ZnS nanocrytals þmagnetic biochar

Pb by 10 times enhanced surface area Yan et al. (2015)

Activated by base 2 M KOH AsV from 24.5 to 31.0 mg g�1 increased surface area Jin et al. (2014)5 M NaOH Pb, Cd by 2.6e5.8 times improved surface area, cation-exchange

capacity, and thermal stabilityDing et al. (2016)

H. Li et al. / Chemosphere 178 (2017) 466e478 475

additional sorption sites for AsV through electrostatic interactionsbetween negatively charged AsV and positively charged Fe oxides.Compared to the control biochar, AsV sorption capacity of themagnetic biochar doubled (265 vs. 429 mg kg�1). Similarly, g-Fe2O3was loaded onto the biochar surface from peanut hull, which sor-bed 1e2 orders of magnitude higher amount of CrVI compared tothe pristine biochar (Han et al., 2016). Samsuri et al. (2013) studiedthe mechanisms of AsV and AsIII sorption by Fe-coated biocharsfrom empty fruit bunch and rice husk. With Fe coating, themaximum AsIII sorption capacity increased from 19 to 31 mg g�1

while AsV sorption increased from 5.5e7.1 to 15e16 mg g�1. Thepossible sorption mechanism was through As complexation withFe3þ on the biochar. Similar mechanism is also proposed forenhanced metal sorption by biochars modified with Mn minerals(Table 3). However, a different mechanism is proposed forenhanced CrVI and AsV sorption by biochar modified with re-ductants such as zero valent Fe or Na2SO3/FeSO4, which are knownto enhance metal reduction and surface complexation with func-tional groups (Zhou et al., 2014; Pan et al., 2014).

Since surface complexation between metals and functionalgroups such as carboxylic, amino, and hydroxyl groups playsimportant roles in metal sorption, various exogenous functionalgroups have been added to biochar. For example, amino groupswere added to biochars from rice husk and saw dust via poly-ethylenimine modification and/or nitration/reduction (Ma et al.,2014; Yang and Jiang, 2014). Due to increased abundance ofamino groups, biochar’s sorption capacity for CrVI increased by

~10-folds. In addition, due to its richness in amino and hydroxylgroups, chitosan has been used tomodify biochar. The CrVI sorptioncapacity of modified biochar increased from 30 to 120 mg g�1,which translated to increased CrVI removal from 27 to 93% (Zhanget al., 2015c; Huang et al., 2016).

Biochar is a porous substance, with surface area significantlyinfluencing its metal sorption ability. Therefore, increasing its sur-face area by incorporating nano-particles enhances its capacity formetal sorption. For example, Yan et al. (2015) synthesized magneticbiochar/ZnS composites by deposing ZnS nanocrystals onto mag-netic biochar. The biochar showed a maximum sorption capacityfor Pb up to 368mg g-1, 10 times higher than that of control biochar.Similarly, Gan et al. (2015) prepared Zn-biochar nanocompositesfrom sugarcane bagasse, showing increased CrVI removal efficiencyof 1.2e2.0 times than that of pristine biochar. Recently, a novelgraphene/biochar composite was synthesized for Hg removal viaslow pyrolysis of graphene-pretreated wheat straw (Tang et al.,2015). Graphene was coated on biochar surface via p-p in-teractions, with loading graphene at 1% resulting in increase in BETsurface area and acidic functional groups from 4.5 m2 g�1 and0.3 mmol g�1 to 17.3 m2 g�1 and 0.5 mmol g�1, increasing Hgsorption capacity from 0.77 to 0.85 mg g�1 and enhancing Hgremoval efficiency from 70% to 80% from a solution containing0.4 mg L�1 Hg.

In addition to incorporating nanoparticles, activation or modi-fication with alkali solution such as NaOH and KOH can also in-crease biochar’s surface area, leading to enhanced metal sorption

H. Li et al. / Chemosphere 178 (2017) 466e478476

capacity. For example, following KOH activation, biochar frommunicipal solid wastes showed enhanced AsV sorption from 25 to31 mg g�1, primarily due to increased surface area from 29 to49 m2 g�1 (Jin et al., 2014). Similarly, NaOH-modified biochar fromhickory wood exhibited 2.6e5.8 times higher sorption capacity formetals than the pristine biochar (Ding et al., 2016).

5. Future research directions

Biochar has potential for metal sorption and has receivedincreasing attention during the past decade. However, studies aremostly at a lab scale, focusing on sorption of single metal fromspiked solution. In natural waters, different heavy metals maycoexist with other pollutants, thereby there is competition forsorption sites on biochar surface between metals and other ions ororganic pollutants. However, by far, few studies have assessed thecompetitive sorption of metals by biochar. Park et al. (2016) usedsesame straw biochar to sorb multi-metals from water, showingthat sorption behaviors of multi-metals (Pb, Cr, Cd, Cu, and Zn)differed from mono-metal sorption due to competition, especiallyfor Cd, which was reduced the most by other metals. Tan et al.(2016) compared the sorption capacity of corn straw biochar foraqueous Hg and/or atrazine, showing that Hg and atrazine inhibitedeach other’s sorption. When phenanthrene and Hg coexisted insolution, Kong et al. (2011) observed direct competitive sorption,each suppressing the other. In addition, humic acids coexist withcontaminants in aqueous environment, possibly influencing metalsorption by biochar. Zhou et al. (2015) showed that humic acidsincreased sorption capacities of Pb and CrVI by biosolid biocharfrom 197 to 233 mmol g�1 and from 688 to 738 mmol g�1. Due to thesorbed humic acids, their functional groups offer additional sites forPb and supply more reducing agent to facilitate the transformationof CrVI to CrIII. Further competitive sorption studies are necessaryto accurately estimate metal sorption capacity of biochar in naturalenvironments.

At present, there is no report of using biochar to remove heavymetals from contaminated wastewater for field application.Contaminated water is more complicated than the simulated waterused by current studies. To make sure the suitability of biochar totreat wastewater, employing physicochemical conditions to simu-late contaminated water or using actual contaminated water forstudies is warranted. In addition, to support field application, futurestudies should address factors related to metal removal efficiency,such as application rate, dosing and recovery approaches, andregeneration and disposal of metal-sorbed biochars. Making bio-char magnetic can help recover biochar following metal sorption.However, recovery of metals sorbed onto biochar and regenerationof biochar are still challenging before its wide acceptance forwastewater treatment.

Economical biochar regeneration can reduce the amount ofbiochar required, therefore decreasing the cost of wastewatertreatment. Wang et al. (2015f) showed that Pb-loaded biochar(eucalypts leaf) could be regenerated using 0.1 M EDTANa2, withhigh Pb desorption efficiency of 84% after 120 min and low loss ofFe. The regenerated magnetic biochar retained high surface areaand pore volume, with little changes in functional groups. A slightdecrease in Pb sorption capacity (52 vs. 42 mg g�1) was observedduring the first regeneration cycle, with no further decrease in thefollowing 5 regeneration cycles (Wang et al., 2015f). Further biocharregeneration studies using different regenerants are necessary todevelop suitable methods to achieve simultaneous metal desorp-tion and retain biochar’s metal sorption ability.

Though biochar has potential for metal removal from water, itsproduction cost is still the main constrain for field application.Making it cost-effective is a key factor for biochar application in

water remediation. Selection of cheap feedstocks for biochar pro-duction, improvement of reuse methods, and enhancing biocharproperties such as surface area and functional groups are alsocritical factors. In summary, it is important to make biochar feasiblefor field application so future study should advance biocharpyrolysis process to explore its full potential to treat metal-contaminated water.

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