Mechanisms of metal sorption by biochars: Biochar characteristics and modifications

Chemosphere 178 (2017) 466e478

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Chemosphere journal homepage: www.elsevier.com/locate/chemosphere

Review

Mechanisms of metal sorption by biochars: Biochar characteristics and modifications 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 b

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

h i g h l i g h t s

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


Biochar properties varied with increasing pyrolysis temperature.
Complexation and electrostatic interaction are important mechanisms for As sorption.
Complexation and reduction are important mechanisms for Cr and Hg sorption.
Cation exchange and precipitation are important mechanisms for Cd and Pb sorption.
Biochar have been modified to enhance its metal sorption capacity.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 10 January 2017 Received in revised form 11 March 2017 Accepted 16 March 2017

Biochar produced by thermal decomposition of biomass under oxygen-limited conditions has received increasing attention as a cost-effective sorbent to treat metal-contaminated waters. However, there is a lack of information on the roles of different sorption mechanisms for different metals and recent development of biochar modification to enhance metal sorption capacity, which is critical for biochar field 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 sorption of As, Cr, Cd, Pb, and Hg by biochar. Biochar properties vary considerably with feedstock material and pyrolysis 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 complexation), Cd and Pb (complexation, cation exchange, and precipitation), and Hg (complexation and reduction). Besides sorption mechanisms, recent advance in modifying biochar by loading with minerals, reductants, organic functional groups, and nanoparticles, and activation with alkali solution to enhance metal sorption capacity is discussed. Future research needs for field application of biochar include competitive sorption mechanisms of co-existing metals, biochar reuse, and cost reduction of biochar production. Published by Elsevier Ltd.

Handling Editor: Patryk Oleszczuk Keywords: Biochar Heavy metal Contaminated water Sorption Functional groups

* Corresponding author. State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, China. ** Corresponding author. State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, China. E-mail addresses: [email protected] (Y. Chen), lqma@ufl.edu (L.Q. Ma). http://dx.doi.org/10.1016/j.chemosphere.2017.03.072 0045-6535/Published by Elsevier Ltd.

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Contents 1. 2.

3.

4. 5.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 467 Characteristics of biochar . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 468 2.1. Surface area and porosity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 468 2.2. pH and surface charge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 468 2.3. Functional groups . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 470 2.4. Mineral composition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 470 Mechanisms of metal sorption by biochar . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 470 3.1. Arsenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 471 3.2. Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 472 3.3. Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 473 3.4. Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 474 3.5. Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 474 Modification of biochar to enhance metal sorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 474 Future research directions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 476 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 476

1. Introduction Heavy metals are ubiquitous in the environment, adversely €rup, 2003). However, various anthroimpacting human health (Ja pogenic activities including mining, smelting, fertilizer and pesticide application, and electronic manufacturing discharge have increased the amount of metal-containing wastewater into the aquatic environment, leading to water contamination with metals. To deal with metal-contaminated water, different methods have been suggested to remove metals from aqueous solution including chemical precipitation, ion exchange, electrochemical treatment, and membrane technologies (Demirbas, 2008). Among the methods, biosorption technique is the most common and costeffective. This is because biosorbents are environmentally friendly and readily available in large quantities, and one of the most popular biosorbents is biochar. Biochar is a carbon-rich, fine-grained, and porous material. It is usually produced by thermal decomposition of biomass under oxygen-limited conditions at temperature <900  C (Lehmann et al., 2006). It has received increasing attention due to its ability to store large amount of carbon, increase crop yield, reduce soil emission of greenhouse 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). More importantly, due to the presence of highly-porous structure and various functional groups (e.g., carboxyl, hydroxyl, and phenolic groups), biochar shows a great affinity for heavy metals (Mohan et al., 2007; Cao et al., 2009; Park et al., 2011). Much research has explored its ability for heavy metal removal from water (Ahmad et al., 2014; Mohan et al., 2014). Biochars are produced from various feedstocks (wood bark, dairy manure, sugar beet tailing, pinewood, and rice husk) at different pyrolysis conditions (temperature, heating transfer rate, and residence time) to sorb metals from 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 with metals in this review. Based on literatures, five mechanisms governing metal sorption from water by biochar have been proposed (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). 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 groups and 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 reduced metal species. The sorption mechanisms and capacity vary considerably with biochar properties and target metals. Recently, researchers reviewed biochar production technologies and metal removal performance (thermodynamics, kinetics, isotherms, capacity, 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). However, most reviews provided sorption mechanisms for metals as a group, lacking a comparison of the main mechanisms for removal of different metals. Since different metals show different species or valence states at different solution pH conditions, the main mechanisms for their sorption are different. Compared to activated carbon, biochar is a promising adsorbent with lower cost for metal removal from water. Metal sorption capacities of biochar are 2.4e147, 19.2e33.4, 0.3e39.1, 3.0e123 mg g1 for Pb, Ni, Cd, and Cr, respectively (Inyanga et al., 2016). However, they are generally lower than that of activated biochar, which are 255 and 91.4 mg g1 for Pb and Cd (Wilson et al., 2006). Therefore, biochars have been modified to enhance their metal sorption capacity by loading biochar with minerals, organic functional groups, reductants, and nanoparticles, and by activating biochar with alkali solution (Mohan et al., 2014). However, so far, there is no review on recent progress on biochar modification except Mohan et al. (2014) who briefly discussed biochar modification by incorporating nanoparticles including magnetic particles and carbon nanotubes. In this review, we aimed to: 1) review the characteristics of biochar to better understand its efficiency in metal sorption, 2) discuss the dominant sorption mechanisms of individual metals by biochar, 3) describe biochar modification to enhance its metal removal from aqueous solutions, and 4) identify research needs and suggest directions for future research. The novelty of the paper is to compare the main mechanisms for removal of different metals by biochar and to review recent progress on biochar modification to enhance metal sorption capacity. This review provides insight into biochar materials and their capability to sorb different heavy metals, which is useful for future research and biochar field application.

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2. Characteristics of biochar Physicochemical properties of biochar significantly influence its ability to sorb metals. Prior to exploring the mechanisms governing metal removal by biochar, its properties need to be well characterized, including surface area, porosity, pH, surface charge, functional groups, and mineral contents. 2.1. Surface area and porosity Surface area and porosity are the major physical properties that influence metal sorption capacity of biochar. When biomass is pyrolyzed, micropores form in biochar due to water loss in dehydration process (Bagreev et al., 2001). Biochar pore size is highly variable and encompasses nano- (<0.9 nm), micro- (<2 nm), and macro-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 pyrolysis temperature. Studies show that elevated temperature generally 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 g1, while surface area increased from 25.4 to 67.6 m2 g1 (Chen et al., 2014). However, it should be noted that, in some cases, biochar

800

produced at high temperature displays lower surface area and porosity. Chun et al. (2004) observed reduced surface area for biochar produced from wheat straw at 700  C compared to that at 600  C (363 vs. 438 m2 g1). Similar result was reported by Jin et al. (2016) who compared biosolids biochar at 600 and 550  C (5.99 vs. 8.45 m2 g1). At high temperature, biochar porous structure maybe destroyed or blocked by tar, leading to decreased surface area. In addition to pyrolysis temperature, the composition of biochar feedstock is also important. For example, the surface area of manure and biosolid biochar (5.4e94.2 m2 g1) is much smaller than that of plant biochar (112e642 m2 g1) such as wheat, oak wood, corn stover, and pine needle (Table 1). Similarly, biosolid biochar shows smaller porosity (0.053e0.068 cm3 g1) than pineneedle biochar (0.076e1.90 cm3 g1) produced at temperature of 500e700  C. Generally, biomass rich in lignin (e.g., bamboo and coconut shell) develops macroporous-structured biochar, while biomass rich in cellulose (e.g., husks) yields a predominantly microporous-structured biochar (Joseph et al., 2007). 2.2. pH and surface charge Similar to surface area and porosity, biochar pH varies with pyrolysis temperature and feedstock (Table 1). Generally, biochar is alkaline with some exceptions depending on feedstock (Table 1). Biochar produced from oak wood at 350 and 600  C was acidic

(A) Surface area

400

2

-1

surface area (m g )

600

200

80 60 40 20 0 0

200

400

600

800

1000

o

Pyrolysis temperature ( C) 14

(B) pH

pH of biochar

12

oak wood corn stover municipal biosolids maize straw swine manure poultry litter manure

10

8

wheat straw pine needle broiler litter manure wastewater sludge conocarpus wastes

6

4 0

200

400

600

800

1000

o

Pyrolysis temperature ( C) Fig. 1. Surface area and pH of biochars produced from various feedstocks at pyrolysis temperature ranging from 100 to 900  C. Data are from Table 1.

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

Wheat straw

Oak wood Corn stover Municipal biosolids

Maize straw

Pine needle

Broiler litter manure Municipal biosolids

Poultry litter manure Swine manure Conocarpus wastes

Wastewater sludge

Oak wood

Wheat straw

Temperature ( C)

Surface area (m2 g1)

300 400 500 600 700 350 600 350 600 400 450 500 550 600 300 450 600 100 200 250 300 400 500 600 700 350 700 500 600 700 800 900 400 600 400 600 200 400 600 800 300 400 500 700 200 400 600 200 400 600

116 189 309 438 363 450 642 293 527 5.49 7.21 7.73 8.45 5.99 1.00 4.00 70.0 0.65 6.22 9.52 19.9 112 236 207 491 59.5 94.2 25.4 20.3 32.2 48.5 67.6 5.4 6.3 5.8 10.6

Porosity (cm3 g1)

0.010 0.010 0.060

0.044 0.095 0.076 0.190 0.000 0.018 0.056 0.053 0.068 0.090 0.099 0.003 0.003 0.008 0.011

pH

4.84 4.91 5.88 6.71 8.46 9.74 9.75 10.5 11.7 9.84 10.5 11.4

Atomic ratio H/C

O/C

0.55

0.29

0.36

0.09

0.15 0.06 0.02 0.07 0.03 1.01 0.87 0.68 0.58 0.43 0.07 0.06 0.03 1.44 1.91 1.08 0.75 0.45 0.33 0.26 0.18

0.05 0.26 0.10 0.37 0.21

0.49 0.35 0.26 0.62 0.48 0.40 0.28 0.17 0.14 0.10 0.10

Content of mineral elements (%) N/C

K

Ca

Mg

Reference P Chun et al. (2004)

0.001 0.001 0.013 0.012 0.120 0.120 0.110 0.110 0.090 0.026 0.023 0.020 0.012 0.013 0.012 0.014 0.013 0.012 0.010 0.011

Nguyen et al. (2010) Nguyen et al. (2010) Jin et al. (2016)

Wang et al. (2015g)

Chen et al. (2008)

Uchimiya et al. (2010) 8.81 9.54 11.1 12.2 12.2 9.50 10.4 10.0 10.4 7.37 9.67 12.2 12.4 5.32 4.87 7.27 12.0 4.60 6.90 9.50 6.11 10.8 11.0

0.48 0.22 0.15 0.03 0.09

0.06 0.04 0.02 0.01 1.05 0.76 0.52 0.30

1.42 0.63 0.26

(4.84e4.91) (Nguyen et al., 2010). Similar low pH at 4.60 was observed for oak wood biochar produced at 200  C, but at 400 and 600  C, the biochar was neutral to alkaline (6.90e9.50) (Zhang et al., 2015a). In addition, biochars from corn stover, wheat straw, and wastewater sludge at low temperature (200e400  C) were also acidic 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 biochar pH and pyrolysis temperature for biochars produced from oak wood (Nguyen et al., 2010; Zhang et al., 2015a), biosolids (Hossain et al., 2011; Chen et al., 2014; Jin et al., 2016), wheat, corn, and maize residues (Nguyen et al., 2010; Wang et al., 2015g; Zhang et al., 2015b), manure (Subedi et al., 2016), and conocarpus wastes (AlWabel et al., 2013). Increasing temperature led to higher ash component, which positively correlated with biochar pH (r ¼ 0.99; Jin et al., 2016), suggesting that ash component is a factor

0.45 0.30 0.30 0.17 0.12

0.41 0.18 0.08 0.06 0.24 0.17 0.02 0.00

0.58 0.20 0.13

0.075 0.064 0.048 0.026 0.029

0.011 0.012 0.009 0.011 0.111 0.102 0.090 0.050 0.014 0.021 0.029 0.018 0.012 0.011

0.85 0.85 0.99 0.93 0.87 3.88 5.88 1.62 3.53 0.04 0.05 0.09 0.12 0.10 0.11 0.18 0.20 0.13 0.38 0.44

5.93 6.27 6.44 6.58 6.96 2.83 3.59 2.03 2.89 4.34 5.18 6.47 6.75 3.47 4.17 4.62 5.35 0.39 1.18 1.39

1.47 1.55 1.64 1.66 1.75 1.73 2.40 1.57 2.13 0.34 0.40 0.48 0.78 0.35 0.43 0.46 0.54 0.04 0.15 0.18

1.82 1.88 2.04 1.93 2.02 1.22 1.54 0.97 1.55 0.08 0.09 0.11 0.13 2.50 2.80 3.30 3.60 0.03 0.06 0.06

Chen et al. (2014)

Subedi et al. (2016) Subedi et al. (2016) Al-Wabel et al. (2013)

Hossain et al. (2011)

Zhang et al. (2015a)

Zhang et al. (2015b)

contributing to biochar high pH. With increasing temperature from 300 to 700  C, contents of total base cations and carbonates in biochar increased, contributing to increased pH (Yuan et al., 2011). In addition, disappearance of acidic functional groups such as eCOOH at high temperature is another contributor. With temperature increasing from 200 to 800  C, basic functional groups on biochar surface produced from conocarpus wastes increased from 0.15 to 3.55 mmol g1, while acidic functional groups decreased from 4.17 to 0.22 mmol g1, consistent with the increased biochar pH from 7.37 to 12.4 (Al-Wabel et al., 2013). Another important property that influences metal sorption by biochar is its surface charge. When biochar is applied to water for metal removal, solution pH strongly influences its surface charge. The point of zero charge (pHPZC) of biochar refers to the solution pH at which its surface net charge is zero. When solution pH is > pHPZC, biochar is negatively charged and binds to metal cations such as Cd2þ, Pb2þ, and Hg2þ. When solution pH is < pHPZC, biochar is

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positively charged and binds metal anions such as HAsO24 and  HCrO 4 . 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 produced from canola, corn, soybean, and peanut straw at 300, 500, and 700  C. At solution pH of 3e7, all biochars were negatively charged. However, compared to biochars produced at 300 and 500  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 negatively charged functional groups on biochars (e.g., eCOOe, eCOH, and eOH) are reduced, resulting in less-negative surface charges and increased pHPZC. 2.3. Functional groups Functional groups such as carboxylic, amino, and hydroxyl groups play important roles in metal sorption. Pyrolysis temperature and biochar feedstock are the two key factors controlling the quantities of functional groups on biochar surface. However, unlike generally increased surface area, porosity, pH, and pHPZC, the abundance of functional groups in biochar decreases with increasing temperature, primarily due to higher degree of carbonization. With increasing temperature, atomic ratios of H/C, O/C, and N/C decrease (Table 1), suggesting decrease in abundances of hydroxyl, carboxylic, and amino groups. FTIR (Fourier Transform Infrared Spectroscopy) spectra have been widely employed to characterize the functional groups on biochar surfaces. The FTIR spectra of functional groups in biochars produced at different temperatures are different. The FTIR spectra of wood and grass biochar changed when pyrolysis temperature increased from 100 to 700  C (Keiluweit et al., 2010). When compared to the feedstock biomass, no significant difference in FTIR spectra was observed at low temperature of 100e200  C, suggesting no change in functional groups. Dehydration of cellulosic and ligneous components in biochar started at 300  C (35003200 cm1, wavenumber) whereas presence of lignin/ cellulose-derived transformation products appeared at 400  C (multiple peaks at 1600700 cm1). An increasing degree of condensation was observed at charring temperature 500  C (intensity loss at 16501500 cm1 relative to that at 885752 cm1) (Keiluweit et al., 2010). During pyrolysis under increasing temperature, most functional groups of lignocellulosic materials are lost. Using FTIR photoacoustic spectroscopy, Yuan et al. (2011) observed decreased intensity of peaks corresponding to carboxylic (eCOOH) and hydroxyl (eOH) groups as temperature increased from 300 to 700  C for biochars from canola, corn, soybean, and peanut straw. Compared to those at 350  C, similar loss of FTIR spectral features was observed 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) spectra have been employed to characterize the functional groups of biochar. Li et al. (2013) investigated the development of functional groups in rice straw and bran biochars produced at 100e800  C using two-dimensional 13C NMR correlation spectroscopy. Biochars from rice straw and bran went through dehydroxylation/dehydrogenation and aromatization process. Generally, with increasing temperature, formation of O-alkylated groups and anomeric O-C-O carbons occured prior to the production of aromatic structures. For biochars produced at temperature <300  C, aliphatic O-alkylated carbons were predominant, which were generally lost in biochars produced at >300  C where aromatic structures were dominant (Li et al., 2013). Similarly, based on NMR spectroscopy, Zhang et al.

(2015b) observed decreased contribution of O-alkyl carbon from 20-54% to 6.9e13% for wheat straw and lignosulfonate biochars as temperature increased from 200 to 600  C, with alkyl carbon being absent 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 responsible for metal sorption from water. They can exchange or precipitate with heavy metals and reduce their availability. Cao et al. (2009) proposed that precipitation of Pb phosphates such as Pb pyromorphite and hydropyromorphite was the main mechanism governing Pb sorption by dairy-manure biochar. During sorption of metal cations (Cd2þ, Cu2þ, Ni2þ, and Pb2þ) onto broiler litter biochar, base cations such as Ca2þ, Mg2þ, Naþ and Kþ were released into the solution from biochar via cation exchange (Uchimiya et al., 2010). Both pyrolysis temperature and feedstock control the amounts of mineral components in biochar (Table 1). Total concentrations of K, Ca, Mg, and P increase with increasing pyrolysis temperature for biochar 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 behave differently from their total concentrations. Generally, during biochar production, water-soluble concentrations of K, Ca, Mg, and P increase when heated at 200  C but decrease beyond that temperature. This is probably due to increased crystallization as evidenced by the formation of whitlockite [(Ca, Mg)3(PO4)2] or incorporation into the silicon structure at pyrolysis temperature of 500  C, which is less soluble (Shinogi, 2004; Cao et al., 2009). Besides temperature, feedstock is also an important factor influencing the concentrations of mineral components in biochar. Generally, P content in oak wood biochar (0.03e0.06%) is much lower than biosolid biochar (1.82e3.60%) (Hossain et al., 2011; Chen et al., 2014; Zhang et al., 2015a). Poultry litter and swine manure biochars generally contain higher K contents (1.6e5.9%) than biochars from other materials (Subedi et al., 2016). In summary, biochar properties vary considerably, mainly depending on pyrolysis temperature and feedstock. However, temperature could have opposite effects on biochar properties, leading to opposite effects on metal sorption. For example, high pyrolysis temperature leads to higher surface area, providing more sites for metal sorption. However, it reduces the amounts of functional groups, which may lead to lower metal sorption via complexation between metals and functional groups. In addition, the impacts of biochar properties on metal sorption is metal dependent as different metals are sorbed via different mechanisms. To obtain biochar with desirable properties for metal removal, understanding the main mechanisms governing metal sorption is the next step. 3. Mechanisms of metal sorption by biochar Table 2 summarizes the sorption capacity and optimum solution pH for metal sorption by biochar. The metal sorption capacity of biochar varies by 1e3 orders of magnitude, ranging from 1 to 200 mg g1 (Table 2). The pH for maximum metal sorption varies with metals, as solution pH significantly influences both metal speciation and surface charge of biochar. Change in solution pH impacts the complexation behavior of functional groups such as carboxyl, hydroxyl, and amino. For example, the ionization of carboxyl group is ~pH 3e4 (Pulido-Novicio et al., 2001). An increase

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in pH makes carboxyl group deprotonated to effectively complex with positively charged metals. For example, with solution pH increasing from 3 to 7, straw biochars were more negatively charged due to enhanced deprotonation of functional groups (Yuan et al., 2011). Five mechanisms have been proposed to govern metal sorption by biochar from aqueous solutions, i.e., complexation, cation exchange, precipitation, electrostatic interactions, and chemical reduction (Fig. 2). However, the role of each of mechanism plays for each metal varies considerably depending on target metals. Previous reviews described sorption mechanisms for metals as a group without comparing the main mechanisms for sorption of different metals. Here, we reviewed the main mechanisms for sorption of As, Cr, Pb, Cd, and Hg. 3.1. Arsenic Arsenic is a carcinogenic metalloid, with concentrations in natural waters varying by several orders of magnitude depending on source and local geochemical environment (Smedley and Kinniburgh, 2002). Arsenic occurs in the environment in several oxidation states (3, 0, þ3, and þ5). In natural water, inorganic arsenate (AsV) and arsenite (AsIII) are predominant, with AsIII being more toxic and mobile than AsV (Manning and Goldberg, 1997). In aerobic environments, AsV is prevalent and exists as 2 H2AsO 4 and HAsO4 at pH 3e11. Under reducing environment, AsIII is dominant and exists as H3AsO03 at pH < 9.2 and H2AsO 3 at pH > 9.2 (Korte and Fernando, 1991; Lenoble et al., 2005). Precipitation and reduction are minor mechanisms for As sorption by biochar. X-ray diffraction (XRD) analyses of pinewood biochar before and after AsV sorption showed no new peaks, suggesting no formation of new minerals and the precipitation mechanism was not important for AsV removal by biochar (Wang et al., 2015e). X-ray photoelectron spectroscopy analysis of a magnetic biochar (produced from water hyacinth with chemical coprecipitation of Fe2þ/Fe3þ) after AsV sorption showed that ~89%

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

of the total As on biochar surface existed as AsV, suggesting little AsV reduction (Zhang et al., 2016). Unlike precipitation and reduction, complexation and electrostatic interactions are important mechanisms for As sorption by biochar. Samsuri et al. (2013) compared AsV and AsIII sorption between two biochars from empty fruit bunch and rice husk. Although the surface area of empty fruit bunch biochar was significantly lower (1.9 m2 g1) than that of rice husk biochar (25.2 m2 g1), its As sorption capacity (5.5 mg g1 AsV and 18.9 mg g1 AsIII) was similar to rice husk biochar (7.1 mg g1 AsV and 19.3 mg g1 AsIII) (Table 2). This could be explained by higher H/C and O/C molar ratios (0.08 and 0.61) for empty fruit bunch

Table 2 Sorption capacity and optimum pH of biochars produced from different feedstocks for metal sorption from aqueous solutions. Metal

Feedstock (pyrolysis temperature)

pH

Sorption capacity (mg g1)

Reference

AsIII

empty fruit bunch rice husk pine wood (400e450  C) pine bark (400e450  C) oak wood (400e450  C) oak bark (400e450  C) empty fruit bunch rice husk peanut straw (400  C) soybean straw (400  C) canola straw (400  C) rice straw (400  C) sugar beet tailing (300  C) coconut coir (produced 250, 350, 500, and 600  C) daily manure (200 and 350  C) oak bark (400e450  C) sludge (550  C) pine wood (400e450  C) pine bark (400e450  C) oak wood (400e450  C) oak bark (400e450  C) hardwood (450  C) corn straw (600  C) hardwood (450  C) corn straw (600  C) almond shell (650  C) Brazilian pepper (300, 450 and 600  C) soybean stalk (700  C)

8.0 8.0 3.5 3.5 3.5 3.5 6.0 6.0 4.0 4.0 4.0 4.0 2.0 3.0 e 5.0 5.0 5.0 5.0 5.0 5.0 5.0 5.0 5.0 5.0 6.0 6.0 7.0

18.9 19.3 1.20 12.2 5.85 7.40 5.50 7.10 25.0 17.2 14.6 14.0 123 31.1, 10.9, 7.90, and 4.10 31.9 and 51.4 5.40 30.9 4.13 3.00 2.62 13.1 6.79 12.5 4.54 11.0 20.0 24.2, 18.8, and 15.1 0.67

Samsuri et al. (2013)

AsV CrIII

CrVI Cd Pb

Cu Zn Ni Hg

Mohan et al. (2007)

Samsuri et al. (2013) Pan et al. (2013)

Dong et al. (2011) Shen et al. (2012) Xu et al. (2013) Mohan et al. (2007) Lu et al. (2012) Mohan et al. (2007)

Chen et al. (2011) Chen et al. (2011) Kılıç et al. (2013) Dong et al. (2013) Kong et al. (2011)

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biochar than rice husk biochar (0.05 and 0.37), suggesting higher amounts of oxygenated functional groups such as carboxyl and hydroxyl groups, which could counterbalance the low As sorption due to low surface area (Samsuri et al., 2013). This suggests that As complexation with functional groups controls As sorption by biochar, which is supported by the shift in FTIR spectra peaks of functional groups including hydroxyl, carboxyl, and C-O ester of alcohols (Samsuri et al., 2013). Following As loading, absorption bands corresponding to hydroxyl group (3386 cm1), CeH groups (2925 cm1), COOe groups (1576 cm1), CH2-groups (1369 cm1), and CeO ester of alcohols, carboxylic acid groups and carboxylic acids (1020e1300 cm1) shifted, while new peaks appeared which were 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 a lower sorption capacity (0.95 mg g1) than that at 300  C (2.84 mg g1) due to loss of oxygenated functional groups. Wang et al. (2015d) measured As sorption by 12 biochars from 4 feedstocks at 3 different temperatures (300, 450, and 600  C), showing similar results that sorbed AsV generally decreased with increasing pyrolysis temperature. These data confirm the importance of functional groups for As sorption by biochar. In addition to complexation, electrostatic interactions is another important 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 AsV sorption of 0.3 mg g1. At solution pH 7, AsV mainly existed as HAsO2 4 while the biochar surface was positively charged, with some functional groups being protonated as solution pH was < pHPZC. AsV oxyanions interact with the positively charged functional groups by electrostatic attraction. At lower solution pH, biochars are more positively charged with higher degree of protonation of functional groups than those at high solution pH, thereby having greater ability to attract AsV oxyanions via electrostatic interactions. Wang et al. (2016b) reported that AsV sorption increased with decreasing solution pH from 9 to 2 by pinewood biochars with pHPZC > 10. Similar pH effects on AsV sorption were observed for Ni/Mn oxide-modified pinewood biochars (Wang et al., 2016a). In short, complexation and electrostatic interactions are important mechanisms for As removal by biochar, with functional groups being the dominant property governing As sorption. Producing biochar at lower pyrolysis temperature using suitable feedstock can help to obtain desirable biochar having higher amounts of functional groups to remove As from water.

with few on CrIII (Wnetrzak et al., 2013; Yang et al., 2013; Pan et al., 2013). However, based on the limited literatures, three mechanisms are responsible for CrIII sorption by biochar: (1) complexation with oxygen-containing functional groups, (2) cation exchange, and (3) electrostatic attraction between positively charged CrIII ions and negatively charged biochar (Fig. 3A). Biochars prepared from crop straws were investigated for CrIII sorption, with sorption capacity following the order of peanut > soybean > canola > rice (0.48, 0.33, 0.28, 0.27 mmol kg1), being consistent with the amounts of oxygen-containing functional groups (1.34, 1.13, 0.80, 0.63 mmol g1). The data suggest that CrIII complexation with functional groups is important for its sorption by biochar (Pan et al., 2013). This was evidenced by peak shifts of functional groups in FTIR spectra for biochar following CrIII sorption, consistent with aromatic C]C ring stretching, phenolic OH region, and aliphatic CeH stretching (Pan et al., 2013). By separating rice straw biochar into biochar colloids (<2 mm) and residues (>2 mm), Qian et al. (2016) observed significantly higher CrIII sorption capacity for biochar colloids than residues, consistent with higher amounts of oxygen functional groups in biochar colloids, implying the role of complexation in CrIII sorption. However, comparing CrIII sorption by pig manure biochar produced at different pyrolysis temperatures, Wnetrzak et al. (2013) observed significantly higher sorption capacity of biochar produced at 600  C than 400  C at solution pH of 4e5 (21e26 vs. 17e19 mg g1). Similarly, Qian et al. (2016) observed significant

3.2. Chromium Chromium (Cr) exists in many valence states ranging from 2 to þ6, with CrIII and CrVI being the major two oxidation states in natural waters (Rai et al., 1989). CrVI is highly soluble and mobile in aqueous solution and is of significant environment concern because of its carcinogenic, mutagenic, and teratogenic behavior in biological systems (Fendorf et al., 2000). CrIII is usually considered as an essential micronutrient for humans, being ~300 times less toxic than CrVI. Although CrIII is considerably less toxic than CrVI, its disposal as a soluble species in natural waters may pose serious health risks because it can be oxidized to CrVI in the environment (Fendorf et al., 2000). Under oxidizing conditions, the principal CrVI 222 species are HCrO 4 , CrO4 and Cr2O7 . Overall, Cr2O7 and HCrO4 2dominate at pH < 6.0 while CrO4 dominates at pH > 6.0 (Richard and Bourg, 1991). In low Eh environments, the main CrIII species are Cr3þ, Cr(OH)2þ, Cr(OH)3(s) and Cr(OH)-4. At pH < 3.6, Cr3þ is the prevalent species. Studies of Cr removal using biochar are mainly focused on CrVI,

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

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increase in CrIII sorption by biochar colloids with increasing temperature from 100 to 400  C. FTIR spectra peak position and intensity showed no significant changes following CrIII sorption by biosolid biochar, suggesting little complexation of CrIII with biochar functional groups (Chen et al., 2015b). They observed release of cations (Ca2þ and Mg2þ) into solution from biochar during CrIII sorption, and the released cations correlated well with sorbed CrIII, suggesting cation exchange is an important sorption mechanism. In addition, during CrIII sorption, slight increase in solution pH was observed (Chen et al., 2015b; Qian et al., 2016), possibly due to release of CaO and MgO from biochar into solution as supported by increased Ca and Mg concentration in solution following CrIII sorption. Sorption of CrIII is pH dependent, with increasing sorption with increasing solution pH at 2.5e5.0 (Pan et al., 2013; Yang et al., 2013). At lower solution pH, higher concentration of Hþ inhibits cation exchange between CrIII and minerals in biochar. However, sorption of CrIII through electrostatic attraction cannot be ruled out. At solution pH of 2.5e5.0, biochars from crop straw were negatively charged based on their negative zeta potential values, while CrIII was mainly positively charged (Cr3þ, CrOH2þ, and Cr(OH)þ 2 ) (Pan et al., 2013). As solution pH increases, biochars become more negatively charged, showing higher electrostatic attraction ability for CrIII. Different from CrIII, two mechanisms have been proposed for CrVI sorption: (1) electrostatic attraction between negatively charged CrVI species and positively charged biochar; and (2) reduction of CrVI to CrIII mainly by oxygen-containing functional groups such as carboxyl and hydroxyl groups and subsequent CrIII complexation with functional groups on biochar (Fig. 3B). Among the mechanisms, CrVI reduction to CrIII followed by CrIII complexation is a major sorption mechanisms for CrVI. Dong et al. (2011) observed that sugar beet tailing biochar was effective in CrVI sorption, which was mainly through CrVI reduction to CrIII and then complexation by hydroxyl and carboxyl groups on biochar, with maximum sorption of 123 mg g1 at pH 2.0 (Table 2). Using coconut coir biochar produced at 250, 350, 500, and 600  C, Shen et al. (2012) observed that at pH 2.0, CrVI sorption decreased sharply from 31.1 to 4.10 mg g1 with increasing temperature from 250 to 600  C, consistent with sharp decrease in acidic functional groups (carboxyl, lactonic, and phenolic) from 1.78 to 0.12 mmol g1. Similar results that low-temperature biochars have higher CrVI sorption capacity have been reported by others (Han et al., 2016; Zhang et al., 2013; Zhou et al., 2016). Using wastewater sludge biochar produced at 300e600  C, Zhang et al. (2013) observed significantly higher CrVI sorption capacity for biochar at 300  C (208 mg g1) compared to that at 400e600  C (36.6e141 mg g1) at pH 2. In addition, longer pyrolysis residence time decreases CrVI sorption capacity. With pyrolysis residence time increasing from 1 to 2 h, municipal wastewater sludge biochars produced at 400e600  C showed decreased CrVI sorption from 69.0e118 mg g1 to 19.6e45.2 mg g1 due to loss of functional groups (Zhang et al., 2013). The data confirm that CrVI reduction followed by CrIII complexation with functional groups on biochar is the main mechanism for CrVI sorption. Reduction of CrVI to CrIII has been confirmed based on spectra analysis of biochar surface by X-ray photoelectron spectroscopy where CrIII and CrVI co-existed, with CrIII being the dominant (93%) species of sorbed Cr (Dong et al., 2011; Zhang et al., 2013). Mohan et al. (2011) further studied CrVI sorption ability of oak wood and oak bark biochars, with wood containing higher amount of lignin, celluloses and hemicelluloses. When oak wood was subjected to pyrolysis, the biochar contained substantial amount of oxygen (8e12%). Oxygen-containing compounds including byproducts catechol, substituted catechol, unsaturated

473

anhydrosugars, and diols generally are effective in reducing CrVI to CrIII (Mohan et al., 2011; Ahmad et al., 2014). Furthermore, the reduction mechanisms can be separated into direct and indirect pathways. The direct reduction occurs in the aqueous phase by solubilized biochar components in aqueous solution (Dong et al., 2014). The indirect reduction occurs on solid biochar surface where CrVI is bound to biochar surface due to electrostatic attraction prior to its reduction to CrIII (Zhou et al., 2016). In short, unlike other metals, biochar enhances CrVI sorption via reduction of toxic CrVI to less toxic CrIII (Fig. 3B). Using biochar produced at low temperature with high amounts of functional groups can facilitate CrVI removal. 3.3. Cadmium Divalent metal cations such as Cd and Pb have a strong tendency to be hydrated in aqueous solution, which is pH dependent. Due to their similarity in aqueous solution, divalent metals share similar sorption mechanisms, i.e., cation exchange, surface complexation, precipitation, and electrostatic interactions. Among divalent metals, sorption of Cd and Pb by biochar has been studied the most. Cadmium has little tendency to be hydrolyzed at pH < 8.0, but at pH > 11.0, all Cd exists as hydrxo-complex. In natural aquatic solution, Cd2þ is the predominant species. Mechanisms for Cd sorption by honey mesquite, cordgrass, and loblolly pine biochars from aqueous solution were reported by Harvey et al. (2011). These plant biochars were grouped into two groups based on their cation exchange capacity: low and high. For the high capacity group, cation exchange was the predominant mechanism for Cd sorption. Flow calorimetry experiment was used to study the behavior of K and Cd to replace Na-saturated biochar. The shape and duration of heat signals of Cd exchange for Na was similar to that of K exchange for Na based on flow calorimetry, indicating that cation exchange is the dominant mechanism (Harvey et al., 2011). Zhang et al. (2015d) showed that released cations (sum of K, Ca, Na, and Mg) from water hyacinth biochar was almost equal to the amount of sorbed Cd, suggesting the dominant role of cation exchange in Cd sorption by biochar. Trakal et al. (2014) tested Cd sorption ability of biochars from waste agro-materials (nut shells, plum stones, wheat straws, grape stalks, and grape husks), showing that grape stalk biochar was more effective in Cd sorption than plum stone biochar (0.45 vs. 0.04 mmol kg1), consistent with its higher cation exchange capacity than plum stone biochar (402 vs. 121 mmol kg1), again suggesting cation exchange as the predominant mechanism. Comparison of FTIR spectra of biochar before and after Cd sorption showed insignificant shift in peaks of carboxyl groups following Cd sorption (Trakal et al., 2014; Chen et al., 2015a), suggesting that Cd complexation with carboxyl groups is a minor mechanism in Cd sorption. However, for honey mesquite, cordgrass, and loblolly pine biochars with low cation exchange capacity, calorimetric heat signals were almost three times higher than that for high capacity group (Harvey et al., 2011), which was inconsistent with cation exchange mechanism. Although cation exchange cannot be ruled out, there is strong evidence that complexation with oxygenated functional groups and electron-rich domains on graphene-like structures is an important mechanism for Cd sorption by biochar with low cation exchange capacity. Due to relatively high soluble concentrations of carbonate and phosphate in dairy manure biochar, precipitation was suggested as the main mechanism for Cd sorption (Xu et al., 2013). With increase in temperature from 200 to 350  C, sorption capacity of Cd increased from 31.9 to 51.4 mg g1 mainly due to increase in minerals especially soluble CO23 in biochar (2.52 vs. 2.94%) (Table 2). Visual MINTEQ modeling coupled with FTIR experiment showed that 88% of Cd sorbed onto dairy manure biochar produced

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at 350  C was attributed to precipitation of metal-phosphate and -carbonate, with 12% of Cd sorption being from Cd-p bonding (Xu et al., 2013). While <3% of the precipitation resulted from metal phosphate due to low soluble P in biochar produced at 350  C, carbonate precipitates accounted for up to 98% of Cd being precipitated. For biochar produced at 200  C, precipitation accounted for 100% sorption of Cd, with 22% from phosphate precipitates due to relatively high content of soluble P with carbonate precipitates accounting for 78% (Xu et al., 2013). X-ray diffraction analyses confirmed the formation of Cd carbonate and phosphate minerals in water hyacinth biochar following Cd sorption (Zhang et al., 2015d). In addition, based on FTIR spectra, Trakal et al. (2014) showed that the peaks of CO23 in high ash content biochar from grape stalks and husks shifted following Cd sorption probably due to surface precipitation of Cd carbonates. 3.4. Lead In natural aquatic system, Pb is present mainly as Pb2þ and Pb(OH)þ at pH < 5.5, Pb(OH)2 at pH 5.5e12.5 and Pb(OH)24 at pH > 12.5 (Ucun et al., 2003). Various biochars including biosolids, dairy manure, oak wood, oak bark, and bagasse have been tested for their ability to sorb Pb from aqueous solutions (Cao et al., 2009; Mohan et al., 2007; Ding et al., 2014). Mechanisms governing Pb sorption 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 Mg is the main one contributing to Pb sorption by biosolid biochar, which explained 40e52% of the Pb sorption at pH 2e5, while exchange with K and Na contributed 4.8e8.5%, with Pb complexation with carboxyl and hydroxyl groups contributing 38e42% (Lu et al., 2012). Cation exchange is also the main mechanism for Pb sorption by oak wood and oak bark biochars, which is supported by the fact that the amount of Pb sorbed onto biochar was similar to that of the cations released (Mohan et al., 2007). Ding et al. (2014) also reported that cation exchange accounted for 62% of Pb sorbed by bagasse biochar produced at 500  C. However, for dairy manure biochar produced at 200, 250 and 350  C, due to the high content of phosphate and carbonate, precipitation as Pb phosphate and Pb carbonate minerals (84e87%) was the predominant mechanism, which was confirmed by chemical speciation, X-ray diffraction, and FTIR spectroscopy data, while complexation with functional groups may contribute to the rest (Cao et al., 2009). By separating biochar (dairy manure and rice straw) into organic and inorganic fractions, Xu et al. (2014) showed that the Pb sorption capacity of organic fraction was only ~ 1 mg g1, while it was >300 mg g1 for inorganic fraction, suggesting limited contribution of Pb complexation with organic functional groups to Pb removal, while cation exchange and Pb precipitation were the dominant mechanisms. For manure biochar, precipitation with phosphate contributed more to Pb sorption than that with carbonate (68 vs. 32%), while the opposite was true for straw biochar (36 vs. 64%). In short, cation exchange, complexation, and precipitation are the three main mechanisms responsible for Cd and Pb sorption by biochars. However, their sorption mechanisms depend on the characteristics of biochar, which is affected by feedstock, pyrolysis temperature, 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 or graphite-like domain was the dominant mechanism for Hg sorption by Brazilian pepper biochar (Dong et al., 2013). With increasing pyrolysis temperature from 300 to 600  C, Brazilian pepper biochar showed decreasing Hg sorption capacity from 24.2 to 15.1 mg g1 at pH 7.0 (Table 2). For biochar produced at 300 and 450  C, 23e31% and 7769% of sorbed Hg was associated with carboxylic and phenolic hydroxyl groups, while for biochar at 600  C, 91% of sorbed Hg was associated with a graphite-like domain on an aromatic structure, with the rest being associated with phenolic hydroxyl groups. The decreased Hg sorption capacity by biochar is probably due to reduced carboxylic and phenolic hydroxyl groups with increasing temperature. Xu et al. (2016) compared Hg sorption by two biochars from bagasse and hickory chips, showing that surface complexation was the most important mechanism. However, different biochar sorbs Hg via different complexation mechanisms. As evidenced by XPS spectra, Hg sorption by bagasse biochar was mainly attributed to the formation of (eCOO)2Hg and (eO)2Hg. Following blocking eCOOH and eOH functional groups using anhydrous methanol and methanol, sorption capacity of Hg by bagasse biochar decreased 18% and 38%. However, the blocking had little effects on Hg sorption by hickory chip biochar, since Hg sorption was mainly resulted from the p electrons of C]C and C]O induced Hg-p binding (Xu et al., 2016). In addition to hydroxyl and carboxylic groups, Hg complexation with thiol groups was also observed based on Hg extended X-ray absorption fine structure analyses (Liu et al., 2016). Thirty-six biochars from different feedstocks were tested for their ability to sorb Hg, with high S biochar showing binding of Hg with S, while Hg was mainly bound to O and Cl in biochars with low S content (Liu et al., 2016). Besides complexation, chemical reduction is responsible for Hg sorption. Kong et al. (2011) investigated Hg sorption by soybean stalk biochar. Besides cation exchange, complexation, and Hg(OH)2 precipitation, they also proposed that Hg2þ was reduced to Hg2Cl2 in presence of Cl, which was then precipitated on biochar surface. However, there was no direct evidence to prove the presence of Hg2Cl2 on biochar. Reduction of Hg2þ to Hgþ by phenol groups or p electrons was observed during the removal of Hg2þ by biochar based on XPS analyses (Xu et al., 2016). Though numerous studies have investigated Hg sorption by activated C, there is no consistent evidence regarding Hg precipitation. Many researchers have attributed Hg sorption on activated C to the presence of Hg(OH)2 due to its existence as an uncharged soluble hydroxide salts, while others have insisted that if Hg precipitates with Cl during sorption, then it is likely to be Hg2Cl2 (Lloyd-Jones et al., 2004). Hence, more studies are needed to prove whether Hg2þ reduction to Hgþ is an important 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 capacity is usually lower compared to other common biosorbents such as activated C. Therefore, recent studies have modified biochar to enhance its metal sorption capacity. For example, efforts have been made to increase its surface area, porosity, pHPZC, and/or functional groups. Approaches to modify biochars include loading with minerals, reductants, organic functional groups, and nanoparticles and activation with alkali solution. Biochar modification includes loading biochar with different minerals such as hematite (g-Fe2O3), magnetite, zero valent Fe, hydrous Mn oxide, calcium oxide, and birnessite (Table 3). The loading can be achieved before, during, or after pyrolysis of the feedstock. Wang et al. (2015e) synthesized a magnetic biochar by pyrolyzing a mixture of hematite mineral and pinewood, thereby incorporating g-Fe2O3 onto biochar surface and serving as

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

modification

Metal

Sorption capacity/removal efficiency

Mechanisms

Reference

Minerals

magnetic biochar of hematite þ pinewood magnetic biochar from iron chloride þ peanut hull iron þ empty fruit bunch

AsV

from 265 to 429 mg kg1

g-Fe2O3 particles as sorption sites

Wang et al. (2015e)

CrVI

1-2 orders higher

AsV, AsIII

iron þ rick husk

AsV, AsIII

Ca/Fe þ rice husk

AsV

amorphous hydrous Mn oxide þ pine wood MnCl2$4H2O or birnessite þ þ pine wood KMnO4 þ hickory wood

Pb

from 5.5 and 18.9 to 15.2 and 31.4 mg g1 from 7.1 and 19.3 to 16.9 and 30.7 mg g1 AsV removal efficiency increased from 25% to 58e95% from 6.4 to 98.9% at pH 5.00

Pb, Cd

Zn þ pine cones

AsIII

increased from 66.1 to 87.6%

Zero valent iron þ bamboo

Pb, CrVI, AsV

Na2SO3/FeSO4 þpeanut straw

CrVI

enhanced Pb, CrVI, and AsV removal efficiency from 23.9, 0.0, and 1.0% to 90e100, 25e40, 20e95% efficient CrVI removal

Reductant

Organic functional groups

Decorated with nano-particles

Activated by base

polyethylenimine þ rice husk chitosan þ magnetic biochar þ Eichhornia crassipes b-cyclodextrinechitosan þ walnut shell Graphene þ wheat straw Zn þ sugarcane bagasse ZnS nanocrytals þ magnetic biochar 2 M KOH 5 M NaOH

Pb, AsV

from 0.2 to 0.59e0.91 for Pb and 2.35 to 4.91e47.1 for AsV by 2.1 and 5.9 times to 153 and 28.1 mg g1

1

CrVI CrVI

from 23.1 to 436 mg g from 30 to 120 mg g1

CrVI

from 27 to 93%

Hg

from 71 to 80%

CrVI Pb

by 1.2e2.0 times by 10 times

AsV Pb, Cd

from 24.5 to 31.0 mg g1 by 2.6e5.8 times

additional sorption sites for AsV through electrostatic interactions between negatively charged AsV and positively charged Fe oxides. Compared to the control biochar, AsV sorption capacity of the magnetic biochar doubled (265 vs. 429 mg kg1). Similarly, g-Fe2O3 was loaded onto the biochar surface from peanut hull, which sorbed 1e2 orders of magnitude higher amount of CrVI compared to the pristine biochar (Han et al., 2016). Samsuri et al. (2013) studied the mechanisms of AsV and AsIII sorption by Fe-coated biochars from empty fruit bunch and rice husk. With Fe coating, the maximum AsIII sorption capacity increased from 19 to 31 mg g1 while AsV sorption increased from 5.5e7.1 to 15e16 mg g1. The possible sorption mechanism was through As complexation with Fe3þ on the biochar. Similar mechanism is also proposed for enhanced metal sorption by biochars modified with Mn minerals (Table 3). However, a different mechanism is proposed for enhanced CrVI and AsV sorption by biochar modified with reductants such as zero valent Fe or Na2SO3/FeSO4, which are known to enhance metal reduction and surface complexation with functional groups (Zhou et al., 2014; Pan et al., 2014). Since surface complexation between metals and functional groups such as carboxylic, amino, and hydroxyl groups plays important roles in metal sorption, various exogenous functional groups have been added to biochar. For example, amino groups were added to biochars from rice husk and saw dust via polyethylenimine modification and/or nitration/reduction (Ma et al., 2014; Yang and Jiang, 2014). Due to increased abundance of amino groups, biochar’s sorption capacity for CrVI increased by

via electrostatic interactions g-Fe2O3 particles as sorption sites via electrostatic interactions As complexation with FeOH2þ and FeOHþ 2 groups As complexation with FeOH2þ and FeOHþ 2 groups metal precipitation and electrostatic interactions increased surface hydroxyls and decreased pHPZC strong AsV and Pb affinity for birnessite particles more surface oxygen-containing functional groups and larger surface area. increased surface hydroxyls and decreased pHPZC metal reduction and surface sorption

Han et al. (2016) Samsuri et al. (2013)

Agrafioti et al. (2014) Wang et al. (2015b) Wang et al. (2015c) Wang et al. (2015a)

Van Vinh et al. (2015) Zhou et al. (2014)

enhanced CrVI reduction and CrIII complexation increased amino groups Increased functional groups

Pan et al. (2014)

enhanced surface area, porosity and thermal stability larger surface area, more functional groups, greater thermal stability increased surface area and pore volume enhanced surface area

Huang et al. (2016)

increased surface area improved surface area, cation-exchange capacity, and thermal stability

Ma et al. (2014) Zhang et al. (2015c)

Tang et al. (2015) Gan et al. (2015) Yan et al. (2015) Jin et al. (2014) Ding et al. (2016)

~10-folds. In addition, due to its richness in amino and hydroxyl groups, chitosan has been used to modify biochar. The CrVI sorption capacity of modified biochar increased from 30 to 120 mg g1, which translated to increased CrVI removal from 27 to 93% (Zhang et al., 2015c; Huang et al., 2016). Biochar is a porous substance, with surface area significantly influencing its metal sorption ability. Therefore, increasing its surface area by incorporating nano-particles enhances its capacity for metal sorption. For example, Yan et al. (2015) synthesized magnetic biochar/ZnS composites by deposing ZnS nanocrystals onto magnetic biochar. The biochar showed a maximum sorption capacity for Pb up to 368 mg g-1, 10 times higher than that of control biochar. Similarly, Gan et al. (2015) prepared Zn-biochar nanocomposites from sugarcane bagasse, showing increased CrVI removal efficiency of 1.2e2.0 times than that of pristine biochar. Recently, a novel graphene/biochar composite was synthesized for Hg removal via slow pyrolysis of graphene-pretreated wheat straw (Tang et al., 2015). Graphene was coated on biochar surface via p-p interactions, with loading graphene at 1% resulting in increase in BET surface area and acidic functional groups from 4.5 m2 g1 and 0.3 mmol g1 to 17.3 m2 g1 and 0.5 mmol g1, increasing Hg sorption capacity from 0.77 to 0.85 mg g1 and enhancing Hg removal efficiency from 70% to 80% from a solution containing 0.4 mg L1 Hg. In addition to incorporating nanoparticles, activation or modification with alkali solution such as NaOH and KOH can also increase biochar’s surface area, leading to enhanced metal sorption

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capacity. For example, following KOH activation, biochar from municipal solid wastes showed enhanced AsV sorption from 25 to 31 mg g1, primarily due to increased surface area from 29 to 49 m2 g1 (Jin et al., 2014). Similarly, NaOH-modified biochar from hickory wood exhibited 2.6e5.8 times higher sorption capacity for metals than the pristine biochar (Ding et al., 2016).

water remediation. Selection of cheap feedstocks for biochar production, improvement of reuse methods, and enhancing biochar properties such as surface area and functional groups are also critical factors. In summary, it is important to make biochar feasible for field application so future study should advance biochar pyrolysis process to explore its full potential to treat metalcontaminated water.

5. Future research directions References Biochar has potential for metal sorption and has received increasing attention during the past decade. However, studies are mostly at a lab scale, focusing on sorption of single metal from spiked solution. In natural waters, different heavy metals may coexist with other pollutants, thereby there is competition for sorption sites on biochar surface between metals and other ions or organic pollutants. However, by far, few studies have assessed the competitive sorption of metals by biochar. Park et al. (2016) used sesame straw biochar to sorb multi-metals from water, showing that sorption behaviors of multi-metals (Pb, Cr, Cd, Cu, and Zn) differed from mono-metal sorption due to competition, especially for Cd, which was reduced the most by other metals. Tan et al. (2016) compared the sorption capacity of corn straw biochar for aqueous Hg and/or atrazine, showing that Hg and atrazine inhibited each other’s sorption. When phenanthrene and Hg coexisted in solution, Kong et al. (2011) observed direct competitive sorption, each suppressing the other. In addition, humic acids coexist with contaminants in aqueous environment, possibly influencing metal sorption by biochar. Zhou et al. (2015) showed that humic acids increased sorption capacities of Pb and CrVI by biosolid biochar from 197 to 233 mmol g1 and from 688 to 738 mmol g1. Due to the sorbed humic acids, their functional groups offer additional sites for Pb and supply more reducing agent to facilitate the transformation of CrVI to CrIII. Further competitive sorption studies are necessary to accurately estimate metal sorption capacity of biochar in natural environments. At present, there is no report of using biochar to remove heavy metals from contaminated wastewater for field application. Contaminated water is more complicated than the simulated water used by current studies. To make sure the suitability of biochar to treat wastewater, employing physicochemical conditions to simulate contaminated water or using actual contaminated water for studies is warranted. In addition, to support field application, future studies should address factors related to metal removal efficiency, such as application rate, dosing and recovery approaches, and regeneration and disposal of metal-sorbed biochars. Making biochar magnetic can help recover biochar following metal sorption. However, recovery of metals sorbed onto biochar and regeneration of biochar are still challenging before its wide acceptance for wastewater treatment. Economical biochar regeneration can reduce the amount of biochar required, therefore decreasing the cost of wastewater treatment. Wang et al. (2015f) showed that Pb-loaded biochar (eucalypts leaf) could be regenerated using 0.1 M EDTANa2, with high Pb desorption efficiency of 84% after 120 min and low loss of Fe. The regenerated magnetic biochar retained high surface area and pore volume, with little changes in functional groups. A slight decrease in Pb sorption capacity (52 vs. 42 mg g1) was observed during the first regeneration cycle, with no further decrease in the following 5 regeneration cycles (Wang et al., 2015f). Further biochar regeneration studies using different regenerants are necessary to develop suitable methods to achieve simultaneous metal desorption and retain biochar’s metal sorption ability. Though biochar has potential for metal removal from water, its production cost is still the main constrain for field application. Making it cost-effective is a key factor for biochar application in

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