Environ Monit Assess (2015) 187:4093 DOI 10.1007/s10661-014-4093-y

Utilization of biochar sorbents for Cd2+, Zn2+, and Cu2+ ions separation from aqueous solutions: comparative study Vladimír Frišták & Martin Pipíška & Juraj Lesný & Gerhard Soja & Wolfgang Friesl-Hanl & Alena Packová

Received: 15 July 2014 / Accepted: 28 October 2014 # Springer International Publishing Switzerland 2014

Abstract The objective of this study was to study the utilization of two different woody-derived biochars for Cd2+, Zn2+, and Cu2+ ions separation from aqueous solutions. Physicochemical characterization confirmed the main differences in sorbent surface area and cationexchange capacity. The maximum cadmium, zinc, and copper sorption capacities were 1.99, 0.97, and 2.50 mg g−1 for biochar (BC) A; 7.80, 2.23, and 3.65 mg g−1 for BC B. Sorption processes can be affected by time and pH. The most of sorbed cadmium and zinc were bound on exchangeable fractions and copper oxidizable fractions. Chemical modification and FT-IR analyses confirmed the crucial roles of hydroxyl and mainly carboxyl functional groups in sorption processes of Cd2+, Zn2+, and Cu2+ ions by BC A and BC B. The

Electronic supplementary material The online version of this article (doi:10.1007/s10661-014-4093-y) contains supplementary material, which is available to authorized users. V. Frišták : M. Pipíška : J. Lesný Department of Ecochemistry and Radioecology, University of SS. Cyril and Methodius, J. Herdu 2, Trnava 917 01, Slovak Republic V. Frišták (*) : G. Soja : W. Friesl-Hanl Department of Health & Environment, Austrian Institute of Technology GmbH, Tulln 3430, Austria e-mail: [email protected] A. Packová Department of Chemistry, University of SS. Cyril and Methodius, J. Herdu 2, Trnava 917 01, Slovak Republic

garden wood rests with leaf mass-derived biochar can be utilized as an effective sorbent for bivalent ions. Keywords Biochar . Feedstock . Sorption . Cd . Zn . Cu

Introduction Heavy metals in wastewater or industrial effluents may pose unforeseen ecotoxicological risks if they escape from the wastewater treatment facilities into the receiving waters. Even if they are withheld and removed from wastewater successfully, they may damage the biology of the treatment plant and require expensive adsorption and cleaning technologies. Production and storage of wastes and hazardous materials may cause the contamination of water, air, or soil recipients with organic and inorganic xenobiotics. Contrary to organic compounds that may be degraded and eventually completely mineralized, heavy metals may migrate but remain largely unmodified (Frišták et al. 2013). Heavy metals in soils are not degraded by natural processes and are persistent for a long time, leading to chronic effects in environment (Bolan et al. 2003). Potentially toxic metals (PTM) pose a risk to public health and ecosystems because of their non-degradability and (eco-)toxicological characteristics (Inyang et al. 2012). Heavy metals enter ecosystems not only from primary industrial sources such as metal plating or mining but also from secondary pollution via wastewater and sewage sludge. PTM effects mainly occur in liquid-phase soluble forms. Therefore, the concentration of PTM in soil solutions is a

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crucial parameter for the assessment of toxic effects. Constituents of the soil matrix like clay minerals and organic matter may reduce the mobility of PTM by various mechanisms, such as adsorption, precipitation, and oxidation-reduction reactions (Tica et al. 2011). Sorption to organic matter and ion exchange are the key processes that control the toxicity, transport, fate, and behavior of heavy metals in soils. Continued deposition of PTM to soil may cause the formation of new mobile heavy metal forms which become leachable and thus available for groundwater and plant roots uptake. For decontamination of heavy metal affected areas, a wide range of techniques can be applied: excavation and land fill-deposition and ex situ or in situ treatment. Conventional dig-and-dump or dig-and-burn technologies are effective but costly. One of the alternative possibilities is the in situ application of soil amendments for immobilization and stabilization of toxic metals in contaminated soil. These methods seem to be more costeffective in comparison with traditional methods due to the possibility of in situ application without further landfill deposition. The design and production of materials for such immobilization measures represent a possibility of using primary and secondary wastes as remediation tools. Friesl-Hanl et al. (2009) and Lee et al. (2012) described the utilization of inorganic soil amendments (zeolites, lime, apatite) for modification of metal mobility and bioavailability but without altering total metal concentration. However, the lack of effective and cheap amendments require the search for new materials or transformation of traditional materials for the development of immobilization agents. The processes of physicochemical transformation or modification of biomass are dependent on a wide range of reaction conditions such as temperature, pressure, and presence of chemical agents or aeration. Kim et al. (2013) showed the effect of different temperatures in the process of pyrolytic decomposition of plant matter on sorption characteristics of the product. Biochar (BC), the carbon-rich product of the slow or fast pyrolysis of biomass such as wood or other organic residues at 400–700 °C, has attracted much scientific attention as an agricultural soil amendment recently (Karer et al. 2013). The application potential of BC in different areas of environmental technologies indicates the manifoldness of this product (Ahmad et al. 2014). BC utilizations extend to soil improvement, climate change mitigation, waste management, water management, energy production, animal husbandry,

Environ Monit Assess (2015) 187:4093

and construction materials (Lehmann and Joseph 2009). In soil management, BC may not only contribute to carbon sequestration but also improve physical soil properties, increase microbial activity, shift microbial consortium compositions, affect levels of microand macronutrients, and thus natural fertility of substrate (Kloss et al. 2014). For effective sorption of inorganic and organic pollutants in soil, carbonaceous materials can be used (Salih et al. 2011). Activated carbon is the most commonly used sorbent from this group and is characterized by enhanced surface area due to thermal or chemical activation. However, scientific studies (Chen et al. 2011; Lu et al. 2012; Xue et al. 2012) have also shown significant ability of BC to act as biosorbent for the removal of different xenobiotics. This utilization is valuable from the point of feedstock availability and costeffectiveness of sorbent preparation. Processing conditions, production methods, and feedstock variability have been shown to affect sorption properties of BC produced in pyrolytic reactors (Ahmad et al. 2014). The content of O-containing carboxyl, phenolic, and hydroxyl functional groups on the surface of BC-based sorbent plays a crucial role in sorption of PTM in soils. Equally, the sorption process can be affected by the presence of non-carbonized fractions in BC matter. These characteristics indicate that BC shows some promise as a new tool in removal technologies of various hazardous substances from wastes or contaminated materials. The effectiveness as a soil or water remediation tool requires the elucidation of the responsible mechanisms that explain how soluble PTM forms are immobilized by BC. The objective of this study was to compare the sorption properties of BC derived from two different woody feedstocks with those of activated carbon for Cd, Zn, and Cu as model substances of bivalent PTM in soils. The effects of contact time, reaction pH, and initial concentration of metals in single-component system on sorption capacity of BC sorbents were studied. FT-IR analyses and chemical modifications were used for understanding the Cd, Zn, and Cu sorption mechanisms. Additionally, the effect of isotridecanol polyglycol ether as a model for nonionic surfactants on sorption capacity of BC was studied. Finally, the fractionation of Cd, Zn, and Cu sorbed onto BC sorbents was studied according to the BCR sequential extraction protocol to examine the desorption properties of previously sorbed metals.

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Biochar samples (denoted as BC A and BC B) were produced in slow pyrolysis process from two different feedstocks: beech wood chips (BC A) and garden green waste residues (BC B). Both materials had maximum dimensions 2 cm×2 cm×2 cm and had been pyrolyzed in a rotary furnace at a highest treatment temperature of 500 °C (residence time 120 min). For ensuring inert and uniform heating conditions, nitrogen was used as flush gas. The biochars were ground and sieved to particles with size <2 mm. Sorbents were pretreated by rinsing in deionized water with conductivity <0.4 μS cm−1 to remove the ash impurities. As a reference material, granulated activated carbon (AC) with particle size <2 mm was used (GUT-W8x30, GUT mbH, Friedberg, Germany). Grain diameter of this bituminous coal-based AC was 0.6–2.4 mm.

centrifuge tubes filled with 20 mL of 0.01 mol L−1 CaCl2 (pH=5.65) and placed on a horizontal shaker for 24 h (200 rpm). The solution of CaCl2 was used to ease phase separation and to keep ionic strength similar to natural soil solutions. Deionized water (<0.4 μS cm−1), spiked with Cd, Zn, or Cu in forms CdCl2, ZnCl2, and CuCl2, was added to each tubes, resulting in concentration range from 10 to 75 mg L−1 and pH=7.0±0.1. Biochar-solution ratio was 1:30. Solution pH changed from 5.65 to 5.50 with increasing concentration of metals. After agitation for 24 h, tubes were centrifuged (38,400×g, 15 min) and supernatant was filtered through a 0.45-μm pore size membrane filter (Schleicher and Schuell, Germany) to remove colloids from the solution. Filters were tested for retaining of metals before use. Concentration of Cd, Zn, and Cu in liquid phase was measured by atomic absorption spectrometry with flame atomization (FAAS, AA 400, PerkinElmer, USA). Sorptions of metal ions were calculated according to Eq. 1:

Physicochemical characterization of sorbents

.
 Qeq ¼ C 0 −C eq  V m

Materials and methods Biochar sorbents production and treatment

The active and potential pH values of BC A and BC B were measured after stirring of biochars with deionized water and 1.0 mol L−1 KCl (ratio 1:2.5) for 1 h and stabilization for 1 h (inoLab pH level 2P, Weilheim, Germany). The electrical conductivities (EC) of BC A and BC B were measured in a 1:10 deionized water extracts after 24-h mixing (inoLab pH level 2P, Weilheim, Germany). The cation-exchange capacities (CEC) of biochars were determined by the BaCl2 method described in our previous paper (Frišták et al. 2013). Surface areas (SA) were estimated by the titration method with NaOH according to Melichová and Hromada (2013). The total C, N, and H contents of biochars were measured by an elemental analyzer (CHNS-O EA 1108, Carlo Erba Instruments, Italy). Initial concentrations of Cd, Zn, and Cu in biochar samples were determined after aqua regia digestion by ICP-MS (PerkinElmer, Elan DRCe 9000).

ð1Þ

where Qeq is the metal uptake (mg g−1), C0 is the initial liquid-phase concentrations of metal (mg L−1), Ceq is the equilibrium liquid-phase concentrations of metal (mg L−1), V is the volume (L), and m is the amount of biochar (g). Adsorption models The obtained equilibrium sorption data were analyzed by mathematical equations of adsorption models with the terms of Langmuir, Freundlich, and DubininKaganer-Radushkevich (DKR). Parameters of adsorption isotherms were calculated by nonlinear regression analysis using the program MicroCal Origin 8.0 Professional (OriginLab Corporation, Northampton, USA). Langmuir is the simplest type of theoretical adsorption model. The isotherm is given by equation: bQmax C eq 1 þ bC eq

Optimization of sorption behavior

Qeq ¼

ð2Þ

The sorption process of Cd, Zn, and Cu from singlecomponent system was studied using a batch equilibration method according to the OECD guideline 106 (OECD 2001) modified by Lair et al. (2006). One gram of BC or AC sorbent was suspended in polypropylene

where Qeq is the amount of sorbed metal at equilibrium (mg g−1), b is the isotherm coefficient characterizing biochar affinity to cadmium ions in solution (L mg−1), Qmax is the maximum metal sorption capacity at saturated biochar binding sites (mg g−1), and Ceq represents

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the metal equilibrium concentration in solution (mg L−1). The Freundlich adsorption model is commonly used to describe the adsorption process on heterogeneous surface. The isotherm is given by equation: . Qeq ¼ KC eq

1

n

ð3Þ

where Qeq is the amount of sorbed metal at equilibrium (mg g−1), K,n are the Freundlich empirical constants characterizing parameters and intensity of sorption process (L g−1), and Ceq is the metal equilibrium concentration in solution (mg L−1). The DKR model is shown to be more general than the Langmuir and Freundlich isotherms. The DKR model has the linear form: lnQeq ¼ lnQm −βε2

ð4Þ

where Qm is the monolayer maximum sorption capacity, β is the activity coefficient related to mean sorption energy, and ε is the Polanyi potential, which is equal to: .   ε ¼ RT ln 1 þ 1 C eq ð5Þ where R is the gas constant (J mol−1 K−1), T is temperature (K). The activity coefficient is related to the sorption energy E, through the following equation: . . 1 2 E ¼ 1 ð−2β Þ ð6Þ

Kinetic study The study of kinetics was performed with batch sorption experiments of Cd, Zn, and Cu ions by BC A, BC B, and AC (33.3 g L−1) at reaction time in the range from 60 to 2880 min. Samples BC A, BC B, and AC were suspended in 20 mL of 0.01 mol L−1 CaCl2 solution (pH=5.65), and tubes were placed on a horizontal shaker for 24 h (200 rpm). Deionized water (<0.4 μS cm−1), spiked with Cd, Zn, or Cu, was added to each tubes, resulting in the required concentration of 60 mg L−1 and pH=7.0±0.1. Biochar-solution ratio was 1:30. After agitation for 24 h, tubes were centrifuged (38,400×g, 15 min) and supernatant was filtered through a 0.45-μm pore size membrane filter (Schleicher and Schuell, Germany) to remove colloids from the solution. Concentrations of Cd, Zn, and Cu in liquid phase were measured

by atomic absorption spectrometry with flame atomization (FAAS). Sorptions of metal ions were calculated according to Eq. 1 as in adsorption studies. The obtained data were evaluated by three kinetic models. Kinetic models For fitting of kinetic data, the models of pseudo-first order, pseudo-second order, and Azizian pseudo-n order were used. The pseudo-first-order equation (Lagergren equation) can be defined as follows: .   dQt dt ¼ k 1 Qeq −Qt ð7Þ in which Qt is the amount of Me2+ ions sorbed at time t (mg g−1), Qeq is its value at equilibrium (mg g−1), and k1 is the rate constant of pseudo-first-order process (min−1). The pseudo-second-order model can be defined as follows: .  2 dQt dt ¼ k 2 Qeq −Qt ð8Þ in which k2 is the rate constant of the pseudo-secondorder process (g mg−1 min−1) and Qt and Qeq have the same meaning that in pseudo-first-order equation. Pseudo-n-order model according to Azizian and Fallah (2010) can be defined as follows: . Qeq n −Qt n dQt dt ¼ k 3 Qn−1

ð9Þ

in which k3 is the rate constant of the pseudo-n-order process (g mg−1 min−1), n is the order of rate equation, and Qt and Qeq have the same meaning as that in pseudo-first-order equation. pH study Experiments for evaluation of effect of pH on sorption process of Cd2+, Zn2+, and Cu2+ ions by BC A, BC B, and AC were carried out by suspending sorbent samples (33.3 g L−1) in CdCl2, ZnCl2, or CuCl2 solutions with pH in the range of 3.0 to 10.0. Firstly, samples BC A, BC B, and AC were suspended in 20 mL of 0.01 mol L−1 CaCl2 solutions with required pH and tubes were placed on a horizontal shaker for 24 h (200 rpm). Subsequently, solutions of Cd, Zn, or Cu were added to each tube, resulting in concentrations of 60 mg L−1 of studied metal. After agitation for 24 h,

Environ Monit Assess (2015) 187:4093

tubes were centrifuged (38,400×g, 15 min) and supernatants were filtered through a 0.45-μm pore size membrane filter (Schleicher and Schuell, Germany) to remove colloids from the solution. Concentration of Cd, Zn, and Cu in liquid phase was measured by atomic absorption spectrometry with flame atomization (FAAS). Sorptions of metal ions were calculated according to Eq. 2 as in previous experiments. Speciation of metals Calculation of the Cd, Zn, and Cu speciation in aqueous solution as a function of total salt concentration, stability constants, and solution pH was performed using the Visual Minteq program (version 2.52). Visual Minteq works with an extensive thermodynamic database for the calculation of metal speciation, solubility, and equilibrium. Data were calculated considering the carbonate system naturally in equilibrium with atmospheric CO2 (p(CO2) = 38.5 Pa). Sequential extraction protocol A modified three-step extraction protocol proposed by the European Community Bureau of Reference (BCR) was used for determination of Cd, Zn, and Cu binding fractions in BC A, BC B, and AC (Rauret et al. 1999). The residual fraction was analyzed as a fourth step. First step: After sorption of Cd, Zn, and Cu, 0.5 g of dried BC A, BC B, and AC was added to the tubes of polypropylene copolymer with 20 mL of 0.11 mol L−1 acetic acid. The extraction mixture was shaken for 16 h at 200 rpm and 22±2 °C. After centrifugation (38,400×g, 15 min) and filtration (membrane filters with a 0.45 μm pore size), liquid phase was decanted into 40 mL and placed at 4± 2 °C. The separated sediment was rinsed with 20 mL of deionized water and the obtained supernatant was also decanted and divided from the sediment. Second step: 20 mL of 0.5 mol L−1 hydroxylamine hydrochloride (pH 2.0±0.1) was added to the sediments from the first step. The extraction mixture was agitated for 16 h at 200 rpm and 22±2 °C. The extraction process described in the previous step was followed. Third step: 5 mL of 8.8 mol L−1 hydrogen peroxide (pH 2.0±0.1) was added in smaller volumes to the

Page 5 of 16, 4093

sediment from the second extraction step. The residue was incubated at 22±2 °C for 1 h with mild shaking. Subsequently, the tubes were heated in a water bath to 85±5 °C for 1 h. The chemical agent was evaporated almost to dryness. This approach was repeated twice. After cooling, 20 mL of 1.0 mol L−1 ammonium acetate (pH 2.0±0.1) was added to rest of BC A, BC B, and AC. The tubes were shaken for 16 h at 200 rpm and 22 °C ±2 °C. The extract was separated as in the previous step. Fourth step: 5 mL of aqua regia mixture was added to the residue from the third step. The digestion process was performed at 95±5 °C to evaporate the chemical agent (4 h). After cooling, 20 mL of 1 mol L−1 nitric acid for extracting the residual metal fractions was used.

Chemical modification Selected chemical agents for modification of BC A, BC B, and AC surfaces were applied according to GardeaTorresdey et al. (1990), Loudon (1984), and Frišták et al. (2013). The esterification of carboxyl functional groups was achieved by shaking sorbents (15 g L−1) in the mixture of concentrated hydrochloric acid and anhydrous methanol (v/v; 1/108) for 6 h at 200 rpm and 22 °C. The general reaction scheme can be described as follows: Hþ

RCOOH þ CH3 OH → RCOOCH3 þ H2 O

ð10Þ

Modified sorbents (BC AF, BC BF, and AC F) were oven dried at 50 °C and used in sorption experiments. Hydroxyl and phenolic functional groups were methylated by suspending the sorbents in formaldehyde (20 g L−1) and shaking for 6 h at 200 rpm and 22 °C. The general scheme is shown as follows: 2ROH þ HCHO→ðROÞ2 CH2 þ H2 O

ð11Þ

The modified sorbents (BC AM, BC BM, and AC M) were oven dried at 50 °C and used in sorption experiments. The effect of isotridecanol polyglycol ether (IDPE) as a model substance for nonionic surfactants was studied by suspending the sorbents (15 g L−1) in 0.05 % solution of IDPE (Neowett-Netzmittel, Kwizda-Agro, Austria) and agitation at 22 °C and 200 rpm for 6 h. Modified

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sorbents (BC AI, BC BI, and AC I) were oven dried at 50 °C and used in sorption experiments. FT-IR The sorbent surface functional groups of BC A, BC B, and AC were measured by infrared spectroscope with Fourier transformation (FT-IR) (Nicolet NEXUS 470, Thermo Scientific, USA) before and after sorption of Cd2+, Zn2+, and Cu2+ ions by KBr technique. The spectra were recorded from 4000 to 400 cm−1 and the ratio of sorbent to KBr was 0.5:100.

Results and discussion The analytical values of merit for the determination of Cd2+, Zn2+, and Cu2+ ion concentrations in aqueous solutions by FAAS were calculated from their calibration curves, and the results are presented in Table 1. The limits of detection (LOD), defined as the concentration giving a signal three times the standard deviation of the blank (signal/noise ratio of 3), were measured for each metal ion in ten independent performances in deionized water as the blank. The limits of quantification (LOQ) as the lowest concentration of Cd2+, Zn2+, and Cu2+ ions that could be quantified in a sample were determined as the concentrations corresponding to a signal/noise ratio of 10. Characterization of BC A, BC B, and AC The main physicochemical properties of studied biochar sorbents compared to activated carbon characteristics are shown in Table 2. Determinations of active and potential pH by adding BC A and BC B into deionized water or KCl solution confirmed the weakly alkaline character of both sorbents. The pyrolysis of cellulosic biomass at temperature >300 °C causes the decreasing of organic acid concentrations and mobilization of soluble alkali salts focused on improving of biochar reaction (Chen et al. 2011). Salinity of BC A aqueous solution was three times lower (0.54 mS cm−1) than salinity of BC B solution (1.67 mS cm−1). EC of BC A was comparable with EC of standard AC solution. Generally, the EC values were related to the temperature of biochar production and feedstock. Kloss et al. (2012) observed higher EC values in straw-derived biochar than in wood-based biochar. Determinations of CEC and SA values indicated the main

Environ Monit Assess (2015) 187:4093 Table 1 LOD and LOQ values for Cd, Zn, and Cu determination by FAAS Cd

Zn

Cu

LOD (mg L−1)

0.002

0.003

0.003

LOQ (mg L−1)

0.007

0.007

0.007

differences between the studied biochars and activated carbon. The values of exchange capacities and surface areas decreased in order BC B>BC A>AC (Table 3). Gaskin et al. (2008) hypothesized that higher concentrations of minerals in feedstock biomass can catalyze the formation of oxygen functional groups on surfaces of pyrolysis products. These functional groups create the negative charges of materials. The SA for BC A and BC B were higher than that of biochar produced from straw at 525 °C by Kloss et al. (2012). On the other hand, the authors showed a higher SA in chars from different feedstock but produced at the same pyrolysis temperature. In poplar- and spruce-derived biochars, values of SA were more than two times higher than for BC A and BC B studied in this paper. The values of sorption surfaces could play a crucial role in sorption processes of inorganic xenobiotics by sorption materials (Frišták et al. 2014). CHN analysis revealed that BC A and BC B sorbents contained comparable levels of carbon, hydrogen, and

Table 2 Physicochemical characteristics of BC A, BC B, and AC BC A

BC B

pH (H2O)

8.78

9.03

8.98

pH (KCl)

8.46

8.77

8.71

−1

EC (mS cm )

AC

0.54

1.67

Ash content (%)

15.20

19.30

Density (kg L−1)

0.36

0.34

CEC(mmol 100 mL−1)

9.83

12.85

3.81

SA (m2 g−1)

27.24

31.54

18.48

C (%)

80.30

79.78

–*

H (%)

1.60

1.59

–*

N (%)

0.40

0.65

0.58 –* 0.49

–*

Cd (mg g−1)

<0.002**

<0.002**

<0.002**

Zn (mg g−1)

0.093

0.095

0.010

0.016

0.021

0.004

−1

Cu (mg g )

*Values were not experimentally determined **Concentration lower than LOD of analytical equipment

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Table 3 Pseudo-first, pseudo-second, and pseudo-n rate constants for sorption process of Cd2+, Zn, and Cu ions by BC A, BC B, and AC Pseudo-first-order rate constants Metal Sorbent Qeq (mg g−1) k1 (min−1) R2 Cd2+

Zn2+

Cu2+

Pseudo-second-order rate constants

Pseudo-nth-order rate constants

Qeq (mg g−1) k2 (g mg−1 min−1) R2

Qeq (mg g−1) kn (g mg−1 min−1) R2

BC A

1.123

0.010

0.950 1.215

0.012

0.991 1.206

0.011

0.999

BC B

1.418

0.018

0.946 1.527

0.017

0.984 1.570

0.017

0.999

AC

1.042

0.002

0.989 1.294

0.001

0.979 1.019

0.002

0.993

BC A

0.955

0.017

0.985 1.018

0.027

0.996 0.999

0.019

0.998

BC B

1.377

0.014

0.961 1.484

0.015

0.995 1.490

0.014

0.999

AC

0.710

0.002

0.977 0.904

0.002

0.964 0.677

0.001

0.998

BC A

1.666

0.018

0.997 1.759

0.018

0.997 1.694

0.017

0.999

BC B

1.759

0.017

0.996 1.850

0.016

0.973 1.743

0.017

0.999

AC

1.407

0.023

0.997 1.473

0.030

0.996 1.442

0.024

0.998

nitrogen. Fuertes et al. (2010) studied the relation between production conditions such as feedstock or pyrolysis temperature and carbon concentration in biochar. The degree of biomass carbonisation could be demonstrated by the level of hydrogen because it was primarily associated with organic matter of plant feedstock. We observed that the carbonization degrees of BC A and BC B were not similar to that for hardwood and corn straw-derived biochars presented by Chen et al. (2011). Elemental analysis showed that both biochars contained <0.002 mg g − 1 Cd, ≤0.095 mg g − 1 Zn, and <0.021 mg g−1 Cu.

Fig. 1 Effect of contact time on sorption capacity of BC A, BC B, and AC for Cd2+ ions. Conditions: sorbent 33.3 g L−1, c0 Cd2+ 60 mg L−1, pH 5.65±0.1; time period: 0, 60, 120, 240, 360, 1440, 2880 min

Sorption process of Cd, Zn, and Cu Kinetic study The processes of Cd2+, Zn2+, and Cu2+ ions separation by BC A, BC B, and AC from single component aqueous solutions were relatively rapid performed. The data obtained from kinetics experiments showed two-steps sorption behavior of BC A and BC B for all studied metals (Figs. 1, 2, and 3). As a first step, the fast metal ions sorption within 120 min of contact time was determined. This step was subsequently followed by the

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Environ Monit Assess (2015) 187:4093 2,0

Fig. 2 Effect of contact time on sorption capacity of BC A, BC B, and AC for Zn2+ ions. Conditions: sorbent 33.3 g L−1, c0 Zn2+ 60 mg L−1, pH 5.65±0.1; time period: 0, 60, 120, 240, 360, 1440, 2880 min

1,8 1,6

BC A BC B AC --- Pseudo-first order - - - Pseudo-second order ..... Pseudo-n order

1,4

-1

Qeq [mg g ]

1,2 1,0 0,8 0,6 0,4 0,2 0,0 0

500

1000

1500

2000

2500

3000

t [min]

uptake of Cd2+, Zn2+, or Cu2+ ions until reaching the reaction equilibrium in 1440 min of sorption process. Sorption process of bivalent ions by bio-sorbents apparently is performed by the first, rapid, and quantitatively predominant step. The rate of this step is due to the abundant availability of active sorption sites. The second step is slower and quantitatively not so significant. Sorption removal of Cd2+ and Zn2+ ions by AC occurred typically as a one-step process with a Langmuir-type of

2,2

Fig. 3 Effect of contact time on sorption capacity of BC A, BC B, and AC for Cu2+ ions. Conditions: sorbent 33.3 g L−1, c0 Cu2+ 60 mg L−1, pH 5.65±0.1; time period: 0, 60, 120, 240, 360, 1440, 2880 min

2,0 1,8

reaction. For Cu2+ ion sorption, the two chars and AC showed very similar behavior (Fig. 3). In order to determine the sorption kinetics of Cd2+, 2+ Zn , or Cu2+ ions by BC A, BC B, and AC, three kinetic models were used: the pseudo-first order (Eq. 7), pseudo-second order (Eq. 8), and pseudo-n order (Eq. 9) were applied to describe the experimental data. The efficiency and applicability of used models were evaluated by coefficients of determination (R2)

BC A BC B AC __ Pseudo-first order - - - Pseudo-second order ..... Pseudo-n order

1,6

-1

Qeq [mg g ]

1,4 1,2 1,0 0,8 0,6 0,4 0,2 0,0 0

500

1000

1500

t [min]

2000

2500

3000

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Our study results confirm that different surface areas of BC A and BC B did not change the character of sorption kinetics while the adsorption capacity increased by introducing the acidic functional groups.

findings as is shown in Table 4. The comparison of obtained kinetic coefficients for separation of Cd2+, Zn2+, or Cu2+ ions by BC A, BC B, and AC showed that Lagergren equation of pseudo-first order described the sorption data with less significance than pseudosecond order for BC A and BC B. On the other hand, pseudo-first order reproduced the sorption data of all three metals for AC better than pseudo-second order. Thus, we confirmed that displacement of alkalineearth ions by cadmium, zinc, and copper ions from the active sorption sites of BC A and BC B being the rate of site loading by ions proportional to square of free sites number (Xue et al. 2012). However, Table 3 shows that pseudo-n order fitted the sorption process of Cd2+, Zn2+, and Cu2+ ions by all three sorbents with highest efficiency as indicated by the R2 values. We observed that experimental and predicted Qeq values obtained by this kinetic model were in most significant agreement. Özer et al. (2007) discussed that instead of assuming first and second order of the reaction, the direct calculation of rate constant and order of the heavy metal sorption process would be a more appropriate approach. The calculated values of Qeq, kn, and reaction order n (data not shown in Table) were consistent with the respective results of the two other models. Fallah and Azizian (2012) showed the applicability and effectiveness of pseudo-n order for heterogeneous surfaces with different sites for sorption or different functional groups for interaction between sorbent and sorbate. Authors demonstrated the effects of sorbent structure and arrangement of micro- and mesopores on the rates of xenobiotic separation processes.

Effect of pH and metal speciation The pH effect is one of the most important parameters for the evaluation of sorption properties. Depending on the element, pH affects the charge of sorbent surfaces, the level of metal ionization, and thus the formation of new metal speciation forms. The study of sorbent behavior under constant sorption conditions but varied initial pH showed the influence of metal species formation in aqueous solution and subsequently sorption capacity of BC A, BC B, and AC for Cd2+, Zn2+, and Cu2+ ions. Based on the speciation analysis of cadmium with the Visual Minteq speciation program, at the initial concentration 60 mg L−1, almost 100 % of cadmium is present in the form of bivalent ions (Cd2+) in the range of pH 3.0–10.0 (Fig. 4). Zinc is present as Zn2+ ions at pH from 3.0 to 8.5 in aqueous solutions (Fig. 5). At pH >7.5, an increase of concentrations of soluble hydroxide forms, mainly Zn(OH)2 and ZnOH+, can be expected. In Fig. 6, the predominance of Cu2+ ions in solution is shown for the pH range to 3.0–7.0. At pH >7.0, the copper ions start to hydrolyze and subsequently may precipitate directly on sorbent surfaces or in solution. The sorption data of cadmium ions (Fig. 4), zinc ions (Fig. 5), and copper ions (Fig. 6) for BC A, BC B, and AC showed the declining competition of metal ions and free protons

Table 4 Langmuir, Freundlich, and DKR equilibrium parameters (±SD) for Cd2+ ions sorption by BC A, BC B, and AC obtained by nonlinear regression analysis Sorbent BC A

BC B

AC

Model

Qmax (mg g−1)

b (L mg−1)

Langmuir

1.99±0.19

0.07±0.015

Freundlich

–

–

DKR

–

–

Langmuir

7.80±0.80

0.08±0.013

Freundlich

–

–

DKR

–

–

Langmuir Freundlich DKR

2.62±0.21 – –

0.63±0.011 – –

K (L g−1) – 0.20±0.012 – – 0.60±0.042

Qm (mg g−1)

1/n – 0.59±0.019 – – 0.87±0.045

–

–

–

–

0.12±0.026 –

0.71±0.074 –

β (mol2 J−2)

R2

–

–

0.998

–

–

0.998

1.36±0.01

−9.55*10−7

0.846

–

–

0.997

–

–

0.995

2.01±0.05

−4.70*10−7

–

–

–

–

1.18±0.01

0.946 0.985 0.981

−7

−6.75*10

0.924

4093, Page 10 of 16

2+

Cd CdCl2 (aq) +

CdCl + CdOH Cd(OH)2(aq)

80

Cd(OH) 80 3Cd(OH) BC A sorption efficiency BC B sorption efficiency AC sorption efficiency 3+

60

60

40

40

20

20

0

0 2

3

4

5

6

7

8

9

10

Qeq (%)

Cdform/Cdtotal*100 (%)

Fig. 4 Effect of pH on the sorption of Cd (60 mg L−1 CdCl2) by BC A, BC B, and AC (33.3 g L−1, d.w.) from singlecomponent solution and theoretical cadmium speciation in reaction solution. Cadmium speciation was calculated using Visual MINTEQ version 3.0 with initial conditions: 60 mg L−1 CdCl2, T=22 °C, pCO2 =38.5 Pa

Environ Monit Assess (2015) 187:4093

11

pH

values, the Qeq of BC A, BC B, and AC for Cd2+, Zn2+, and Cu2+ ions decreased due to the formation of new metal hydroxide forms, phosphates, and carbonates mainly in the case of Zn and Cu. Similar results were observed by Kolodynska et al. (2012). The authors also suggested the precipitation and ion exchange as main processes involved into sorption of metal cations by manure-derived biochar that depends on changes of initial pH of reaction solution.

for unoccupied binding sites with increased solution pH. The release of binding sites on sorbent surfaces causes the improvement of sorption effectiveness for bivalent metal ions (Frišták et al. 2014). The maximum sorption capacity of BCA, BC B, and AC and thus removal efficiency of Cd2+, Zn2+, and Cu2+ ions was reached at pH 7.5 for cadmium and at a pH of about 7.0 for zinc and copper. Consequently, for other sorption experiments, pH 7.0 was used. At higher pH

2+

Zn + ZnCl + ZnOH 100 ZnOH2 (aq)

100

3-

60

60

40

40

20

20

0

0 2

3

4

5

6

7

pH

8

9

10

11

Qeq (%)

ZnOH BC A sorption efficiency BC B sorption efficiency AC 80sorption efficiency

80

Znform/Zntotal*100 (%)

Fig. 5 Effect of pH on the sorption of Zn (60 mg L−1 ZnCl2) by BC A, BC B, and AC (33.3 g L−1, d.w.) from singlecomponent solution and theoretical zinc speciation in reaction solution. Zinc speciation was calculated using Visual MINTEQ version 3.0 with initial conditions: 60 mg L−1 ZnCl2, T= 22 °C, pCO2 =38.5 Pa

Environ Monit Assess (2015) 187:4093

2+

Cu + CuCl + CuOH 3+ Cu2(OH)

100

Cu(OH) 100 2(aq) 2+

Cu2(OH)2

2+

Cu3(OH)4

60

60

40

40

20

20

0

0 2

3

4

5

6

7

8

9

10

Qeq (%)

BC A sorption efficiency 80B sorption efficiency BC AC sorption efficiency

80

Cuform/Cutotal*100 (%)

Fig. 6 Effect of pH on the sorption of Cu (60 mg L−1 CuCl2) by BC A, BC B, and AC (33.3 g L−1, d.w.) from singlecomponent solution and theoretical copper speciation in reaction solution. Copper speciation was calculated using Visual MINTEQ version 3.0 with initial conditions: 60 mg L−1 CuCl2, T=22 °C, pCO2 =38.5 Pa

Page 11 of 16, 4093

11

pH

BC A, BC B, and AC in solution caused minimal differences of initial and equilibrium pH. The applied pretreatments of sorbents with 0.01 mol L−1 CaCl2 avoided the significant pH changes after 24 shaking. Sorption isotherms The comparison of Cd2+, Zn2+, and Cu2+ ions separation by BC A, BC B, and AC was studied in the concentration range of 10–75 mg L−1. Previous kinetic studies have suggested that the sorption removal of metal ions is a longer time process and the sorbents reach equilibrium within 24 h. Sorption data of cadmium, zinc, and copper separation obtained after this equilibrium time, at pH 7, and at desired initial metal concentrations showed adsorption isotherms of typical “L” shape. The sorption behavior could be described with the Langmuir (Eq. 2), Freundlich (Eq. 3), and DKR (Eq. 4) adsorption models. The sorption constants and correlation coefficients for Cd2+, Zn2+, and Cu2+ ions onto studied sorbents are given in Tables 4, 5, and 6. Cadmium removal processes were well fitted by Langmuir and Freundlich models in the case of BC A and BC B (R2 >0.95). For AC, the same trend but with lower coefficient of determination appeared. The calculated maximum sorption capacity (Qmax) of BC B for cadmium ions (7.80 mg g−1) was nearly four times higher than Qmax of BC A (1.99 mg g−1). The sorption capacities of BC A and BC B were higher than those of a biochar produced from fast pyrolysis of wood and bark (Mohan et al. 2007) and comparable to the sorption

capacity of biochar produced at 500 °C from Miscanthus sacchariflorus biomass (Kim et al. 2013). The comparison of cadmium sorption efficiency between biochars and AC showed the significant decrease of AC Qmax for cadmium (2.62 mg g−1) due to differences in activated carbon surface area. This trend also could be observed in further sorption experiments with zinc and copper. Kolodynska et al. (2012) discussed the effect of surface area on the sorption properties of biochars and pointed out that the degree of carbonization was the most important parameter for inorganic species removal. Carbonized fractions preferred the physical surface sorption. On the other hand, the parameter b that is related to the affinity of the binding sites (Frišták et al. 2014) allows the comparison of the sorbent affinities toward the cadmium ions. Sorption affinity increased in the order BC A BC A (0.97 mg g − 1 ) > AC (0.85 mg g−1). Garden wood residue-derived biochar

4093, Page 12 of 16

Environ Monit Assess (2015) 187:4093

Table 5 Langmuir, Freundlich, and DKR equilibrium parameters (±SD) for Zn2+ ions sorption by BC A, BC B, and AC obtained by nonlinear regression analysis Sorbent

Model

Qmax (mg g−1)

b (L mg−1)

BC A

Langmuir

0.97±0.02

0.57±0.01

Freundlich DKR BC B

Langmuir

– 2.23±0.01

– –

–

–

DKR

–

–

0.85±0.05

Freundlich

–

–

DKR

–

–

– 0.31±0.06

–

–

–

–

1.00±0.03

0.36±0.01

Qm (mg g−1)

1/n

– 0.37±0.01

1.02±0.10

Freundlich Langmuir AC

–

K (L g−1)

0.36±0.04

–

–

–

–

0.27±0.01

0.32±0.04

–

showed lower capacity for Zn2+ ion sorption in comparison with Cd2+. Langmuir adsorption model reproduced the zinc sorption data with the highest efficiency (R2 > 0.98). Sorption affinity of sorbents for zinc increased in the order AC0.95). Our analyses showed that maximum sorption capacity for Cu2+ ions decreased in the order BC B (3.65 mg g−1) > BC A (2.50 mg g−1) > AC (2.24 mg g−1). These capacities are in agreement with those of brewers draft-derived biochars (Trakal et al. 2014) and higher than those of softwood-derived biochar (Han et al. 2013), or switchgrass-derived biochar (Han et al. 2013). Sorption affinity of studied

–

β (mol2 J−2)

–

–

–

–

0.84±0.01 – – 1.92±0.11 – – 0.72±0.01

R2 0.994 0.940

−9

−2.52*10 –

0.907 0.990

–

0.975

−1.41*10−9 –

0.936 0.987

–

0.970

−4.89*10−9

0.862

sorbents for copper increased in the order BC B
Table 6 Langmuir, Freundlich, and DKR equilibrium parameters (±SD) for Cu2+ ions sorption by BC A, BC B, and AC obtained by nonlinear regression analysis Sorbent

Model

Qmax (mg g−1)

b (L mg−1)

BC A

Langmuir

2.50±0.09

1.31±0.01

BC B

Freundlich

–

–

DKR

–

–

Langmuir

0.85±0.06

–

–

DKR

–

–

Langmuir AC

3.65±0.01

Freundlich

Freundlich DKR

2.24±0.06 – –

1.16±0.04 – –

K (L g−1) – 1.32±0.06 – – 1.06±0.02 – – 1.13±0.03 –

Qm (mg g−1)

1/n – 0.55±0.01 – – 0.70±0.01 – – 0.53±0.01 –

β (mol2 J−2)

R2

–

–

0.992

–

–

0.991

1.88±0.01

−8.22*10−8

–

–

–

–

2.11±0.06

–

–

–

1.58±0.03

0.993 0.991

−7.59*10−8

–

0.948

0.989 0.998 0.993

−8

−6.08*10

0.917

Environ Monit Assess (2015) 187:4093

Page 13 of 16, 4093

Table 7 Effect of chemical modification of BC A, BC B, and AC by methylation (BC AF, BC BF, AC F), esterification (BC AM, BC BM, AC M), and IDPE (BC AI, BC BI, AC I) application on sorption of Cd2+, Zn2+, and Cu2+ ions Sorbent

Cd Qeq (mg g−1)

Zn Change (%)

Cu

Qeq (mg g−1)

Change (%)

Qeq (mg g−1)

Change (%)

BC ANa

1.20±0.007

BC AF

1.18±0.009

−1.50

0.54±0.012

BC AM

0.50±0.001

−58.06

0.05±0.002

BC AI

1.26±0.001

+5.10

1.15±0.068

BC BNa

1.60±0.012

BC BF

1.54±0.001

−3.38

0.69±0.021

−52.42

1.63±0.02

−4.12

BC BM

1.29±0.009

−19.42

0.19±0.005

−86.71

1.38±0.01

−18.83

BC BI

1.72±0.002

+8.02

1.70±0.094

+18.88

1.72±0.02

+1.18

AC Na

0.96±0.012

AC F

0.61±0.002

−36.39

0.46±0.009

−41.77

1.39±0.01

AC M

0.41±0.001

−57.03

0.13±0.004

−83.54

1.27±0.01

−9.09

AC I

1.01±0.001

+5.53

0.82±0.038

+3.81

1.41±0.01

+0.72

a

1.01±0.041

1.65±0.02 −46.83

1.62±0.02

−1.82

−95.35

1.48±0.01

−10.33

+13.76

1.67±0.01

+1.23

1.43±0.012

1.70±0.01

0.79±0.021

1.40±0.01 −0.71

BC AN, BC BN, and AC N represent unmodified sorbents

cadmium on this type of biochar fraction than BC A. Gu et al. (2013) reported that the strong association of Cd to carbonate may be attributed to the similarity of the ionic radii of Cd (0.097 nm) and Ca (0.099 nm). When calcium ions precipitate with carbonates, the cadmium ions diffuse into the calcite crystal lattice as a camouflaged element and co-precipitates with carbonates. The reducible fraction of cadmium was determined as cadmium associated with Fe/Mn oxides of BC A (0.31 mg g −1 ), BC B (0.33 mg g −1 ), and AC (0.04 mg g−1). The oxidizable fractions of cadmium associated with organic matter and sulphides amounted to 0.03 mg g−1 for BC A, 0.04 mg g−1 for BC B, and 0.01 mg g−1 for AC. The residual fraction of cadmium such as cadmium within crystal structures of primary or secondary minerals quantified after aqua regia digestion of matter was 0.01 mg g−1 for BC A, 0.02 for BC B, and 0.02 for AC. The total recovery range of cadmium from BC A, BC B, and AC was 82.4–98.7 %. For sorbed zinc by BC A, BC B, and AC, a similar fractionation trend was found. The majority of bound zinc was removable from BC A (0.65 mg g−1), BC B (1.15 mg g−1), and AC (0.62 mg g−1) by 0.11 mol L−1 acetic acid. The zinc associated with Fe/Mn oxides of BC A (0.27 mg g−1), BC B (0.25 mg g−1), and AC (0.09 mg g−1) was determined. Zinc associated with organic matter and sulphides was 0.02 mg g−1 for BC

A, 0.03 mg g−1 for BC B, and 0.03 mg g−1 for AC. The residual fraction of zinc after aqua regia digestion of matter was 0.02 mg g−1 for BC A, 0.03 mg g−1 for BC B, and 0.04 mg g−1 for AC. The total recovery range of zinc from BC A, BC B, and AC was 91.1–99.3 %. Cu speciations in BC A, BC B, and AC were dominant in the fraction associated with organic matter and sulphides (61.8–72.4 %). For the other fractions, the desorption of copper followed the order of the chemical forms residual>acid soluble>reducible. Copper fractions of sorbents with high pH values are predominantly associated with oxidizable and reducible fractions (Gu et al. 2013). The total recovery range of copper was 96.5–104.4 %. Mechanisms of Cd, Zn, and Cu sorption process Chemical modification of sorbents Chemical modifications of BC A, BC B, and AC were carried out to elucidate the contributions of the carboxyl and hydroxyl functional groups on the biochar surface for the Cd2+, Zn2+, and Cu2+ ion sorption and the sorption behavior of biochar. The effects of the chemical modification of the studied sorbents are shown in Table 7. The blocking of carboxyl functional groups caused the most dramatic decrease of BC A, BC B,

4093, Page 14 of 16

and AC sorption capacity for all three studied metals. Esterification of –COOH decreased the Qeq of BC A by about 58.1 % for cadmium, 95.3 % for zinc, and 10.3 % for copper. In the case of BC B, esterification caused a decrease of Qeq by about 19.4 % for cadmium, 86.7 % for zinc, and 18.8 % for copper. The same trend we observed with AC. Similar decreases in the metal separation from aqueous solutions has been reported by Iqbal et al. (2009) when the carboxyl groups of algaederived sorbents were blocked. On the other hand, the methylation of hydroxyl and phenolic functional groups of BC A, BC B, and AC affected the sorption capacity less significantly. Blocking of –OH decreased the Qeq of BC A only by about 1.5 % for cadmium, 46.8 % for zinc, and 1.8 % for copper. The effect of formaldehyde on BC B revealed a decrease of Qeq by about 3.4 % for cadmium, 52.4 % for zinc, and 4.2 % for copper. Sorption capacity of AC was lower by about 36.3 % for cadmium, 41.8 % for zinc, and 0.7 % for copper. This indicates that Cd2+, Zn2+, and Cu2+ ions were sorbed by inner-sphere complexation with the free carboxyl functional groups. The formation of Cd, Zn, and Cu complexes with free phenolic and hydroxyl functional groups apparently contributed less to the separation process. However, we observed that methylation reduced the sorption capacity of BC B more than BC A focused on higher concentration of hydroxyl groups on sorbent surfaces. Lu et al. (2012) discussed the effects of methylation and esterification on sludgederived biochar and pointed at the crucial role of carboxyl groups that were supposed to be involved in the sorption process of Pb2+ ions. Chemical modification of BC A, BC B, and AC by IDPE simulated the effect of nonionic surfactants, included in some crop fertilizers, on sorption properties of biochars. After chemical modification, the Qeq values of all three sorbents increased in comparison with those of non-treated materials. Frišták et al. (2013) discussed the impact of sodium dodecyl sulphate on the sorption capacity of dried activated sewage sludge. The authors explained the improvement of Qeq values for Co2+ ions by the removal of lipid rests from structure. Nonionic surfactants may have the potential to extract residual pyrolysis condensation products from biochar surfaces. FT-IR analyses FT-IR spectra of BC A, BC B, and AC before sorption of metal cations indicated –COOH and –OH groups

Environ Monit Assess (2015) 187:4093

were abundantly present. As shown by chemical blocking of sorbent functional groups, these groups can be involved as proton donors into sorption processes of Cd, Zn, and Cu ions. Spectral analyses of BC A revealed the changes in absorption intensity and shifts of bands at 3418–3421 cm−1 after Cd2+ sorption, at 3418–3423 cm−1 after Zn2+ sorption, and at 3418– 3422 cm−1 after Cu2+ sorption. Analyses of BC B spectra showed shifts at σ 3415–3412 cm−1 after Cd2+ sorption, at σ 3415–3411 cm−1 after Zn2+ sorption, and at σ 3415–3424 cm−1 after Cu2+ sorption. FT-IR analyses of AC showed the band shifts at 3419 to 3423 cm−1 after Cd2+ sorption and to σ 3422 cm−1 after Zn2+ and Cu2+ sorption. These shifts pointed to the complexation of metal ions with hydroxyl functional groups on the biochar surfaces (Frišták et al. 2013). Comparison of FT-IR spectra before and after sorption of cadmium, zinc, and copper ions showed also the significant shifts of wavenumber responsible for stretching of C=O and C=C groups of aromates at 1579 and 1445 cm−1 for BC A, at 1581 and 1437 cm−1 for BC B and also for AC. The absorption band of BC A at 1579 cm−1 was shifted to 1575 cm−1 after cadmium uptake, to 1576 cm−1 after zinc uptake, and to 1587 cm−1 after uptake of copper. The absorption band of BC B at 1581 cm−1 was shifted to 1576 cm−1 after cadmium and zinc uptake and to 1572 cm−1 after uptake of copper. FT-IR of AC showed also shifts from 1581 to 1579 cm−1 after cadmium and zinc sorption and to 1571 cm−1 after copper sorption. The changes of σ characteristic for stretching vibration of carboxylate functional groups might have been caused by the coordination of bivalent metal ions in forms like –COOMe (Chun et al. 2004). For C-H bending, no changes in the spectra of BC A, BC B, and AC after sorption process of all three ions were found.

Conclusions This study investigated the utilization of two different biochars for the separation of Cd2+, Zn2+, and Cu2+ ions from aqueous solutions. The sorptions were affected by pH value, contact time, initial metal concentration, and surfactant modification. The sorption efficiency of wood-based biochar increased in the order Zn>Cd> Cu, whereas biochar from green garden residues was sorbed in the order Zn>Cu>Cd. The analysis of mechanisms of separation processes of metal ions indicated a major role of ion-exchange processes involving

Environ Monit Assess (2015) 187:4093

carboxyl groups. The biochar from garden residues could be utilized as a more effective sorbent for bivalent ions separation than a classical wood-based biochar.

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