Adsorption of heavy metals from aqueous solution by UV-mutant Bacillus subtilis loaded on biochars derived from different stock materials

Ecotoxicology and Environmental Safety 148 (2018) 285–292

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Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv

Adsorption of heavy metals from aqueous solution by UV-mutant Bacillus subtilis loaded on biochars derived from different stock materials

MARK

⁎

Ting Wanga,b, Hongwen Suna, , Xinhao Rena,c, Bing Lia, Hongjun Maob a b c

MOE Key Laboratory of Pollution Processes and Environmental Criteria, College of Environmental Science and Engineering, Nankai University, Tianjin 300071, China Centre for Urban Transport Emission Research, College of Environmental Science and Engineering, Nankai University, Tianjin 300071, China School of Environmental Science and Engineering, Shaanxi University of Science and Technology, Xi’an 710021, China

A R T I C L E I N F O

A B S T R A C T

Keywords: Adsorption Heavy metals Bacterium Biochar Carrier Competitive adsorption

Two kinds of biochars, one derived from corn straw (CBC) and one from pig manure (PBC), were used as the carriers of a bacterium (B38) to adsorb heavy metals in solution. CBC exhibited high affinity to Hg(II), while PBC showed large adsorption capacity of Pb(II). After loading with B38, the sorption capacity of the co-sorbents were enhanced for Pb(II), but weakened for Hg(II). In a binary system, the overall adsorption capacity to Hg-Pb (CBC+B38, 136.7 mg/g; PBC+B38, 181.3 mg/g) on co-sorbents was equal to the sum of the single-component values for Hg(II) and Pb(II). Electrostatic interactions and precipitation are the major mechanisms in the adsorption of Hg(II). In contrast, cation-π interactions and precipitation were involved in the sorption process of Pb (II). Moreover, the sorption sites of Hg(II) and Pb(II) partially overlapped on the biochar surface, but were different on co-sorbents. Hence, the co-sorbents have an advantage over the biochar alone in the removal of heavy metal mixtures.

1. Introduction Heavy metals are persistent environmental contaminants because they cannot be degraded or destroyed (Coral et al., 2005; Lu et al., 2017). Mercury (Hg) and lead (Pb) has been classified as priority inorganic pollutants, exposure to which can cause a wide range of adverse health consequences at very low levels (Jazi et al., 2014; Arshadi, 2015). Biosorption using microbial biomass as the adsorbent has emerged as a potential alternative technique to the existing methods for heavy metal removal (Öztürk, 2007). An encountered problem is the toxicity of heavy metal contaminants, which presents a stress on growth performance of the microflora. Therefore, there is an acquirement of microbial species with high heavy metal-tolerance. Another limitation is the difficulty of separating microbial biomass from solution after adsorption. Thus, the selection of an optimal matrix as the carrier of microflora is vital to the application of biosorption. Various polymeric materials such as polysulfone, alginate, polyacrylamide, and polyvinyl alcohol have been used as carriers for bacterial immobilization (Hu and Reeves, 1997; Al-Hakawati and Banks, 2000; Arıca et al., 2003). However, most of the polymeric materials are not suitable for practical application because of their possible toxicity to bacteria, high cost, short service life, requirement for recycling, and environmentally destructive nature (Al-Hakawati and Banks, 2000; Kim et al., 2014).

⁎

Corresponding author. E-mail address: [email protected] (H. Sun).

http://dx.doi.org/10.1016/j.ecoenv.2017.10.039 Received 15 May 2017; Received in revised form 3 October 2017; Accepted 16 October 2017 0147-6513/ © 2017 Elsevier Inc. All rights reserved.

Biochar, a partially carbonized material produced from the oxygenlimited pyrolysis of biomass at low temperatures (< 700 °C) (Azargohar and Dalai, 2006). Biochar contains plentiful micro- and nano-pores with aromatic structure, which make it strong adsorbent for various contaminants, including heavy metals (Harvey et al., 2011; Chen et al., 2014). In addition to the porous moiety, biochar has a large amount of adsorption sites on the surface, such as carboxyl and hydroxyl groups which show strong chelating capacity for heavy metals (Zhang et al., 2013; Trakal et al., 2014). Cao et al. (2009) evaluated the ability of dairy-manure-derived biochar to sorb Pb(II). It was revealed that lead precipitation as β-Pb9(PO4)6 (84–87%) and surface sorption (13–16%) via the coordination of Pb d-electrons to C˭C (π-electrons) and -O-Pb bonds were the main sorption mechanisms. The atomic mass of Hg is similar with that of Pb, and both Hg and Pb are belong to the top five most toxic heavy metals. However, few data are available for Hg(II) sorption mechanisms by biochar, and for competitive sorption in the presence of Pb(II). Besides the strong sorption capacity to chemical contaminants, the presence and size distribution of the pores in biochar provides a suitable habitat for many microorganisms by protecting them from predation and desiccation and by meeting their diverse needs of carbon (C), energy and mineral nutrient (Saito and Muramoto, 2002; Warnock et al., 2007). Thus, biochar is thought to be able to provide a good

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2.3. Characterization of biochar and co-sorbent

habitat for microorganisms. Thies and Rillig (2009) detected the existence of a range of microbial communities within biochar pores. Hence, biochar could be a good carrier for microorganisms. However, there is little report on the effect of using biochar as a carrier of microorganisms for heavy metal immobilization. The structure and properties of biochars vary with their raw materials. Biochars derived from plant residues commonly contain low amounts of inorganic moiety (ash) (Brewer et al., 2011; Spokas et al., 2011) and usually do not provide macro- and micro-nutrients in sufficient quantities (Glaser and Birk, 2012). Biochars derived from livestock manures may have quite different compositions compared to those from plant residues. One primary difference is that biochars derived from livestock manures usually contain high ash contents (Cao et al., 2010; Zhang et al., 2013) and hence are richer in essential nutrients than plant-derived biochars (Sarkhot et al., 2012). In a former study, we obtained a mutant genotype (B38) from the wild-type Bacillus subtilis using UV irradiation (Jiang et al., 2009). The mutant genotype exhibited 4-fold greater resistance to cadmium (Cd) than the original wild-type species in aqueous solution and has high affinities for Cd, chromium (Cr), Hg, and Pb ions in solution (Wang and Sun, 2013). In the present study, two kinds of biochar, derived from plant residue or livestock manure, were loaded with B38 to form the cosorbents to check their potential as carriers for microorganism to remove heavy metals from solution. The adsorption behavior of two heavy metal cations, Hg(II) and Pb(II), on the biochars and the cosorbents with B38 were studied in single- and bi-solute solutions.

The bulk organic elemental compositions (C, H, and N) of biochar sample were determined by an element analyzer (Elementar Vario EL, Germany). The ash content of biochar samples was measured by the residual weight after heating at 750 °C for 6 h (Keiluweit et al., 2010), and the O content was calculated by the mass difference. The surface morphology of biochar sample was investigated by scanning electron microscopy (SEM, QUANTA 200, USA). The surface elemental compositions of the biochars and the cosorbents were quantified using an X-ray photoelectron spectrometer (XPS) (Kratos Axis Ultra DLD, UK). Fourier transform infrared spectroscopy (FTIR) spectra of the biochars and the co-sorbents were recorded between 4000 and 400 cm−1 wavenumbers using an FTIR spectrometer (Tensor 27, Bruker, USA). 2.4. Preparation of the metal ion solutions The metal ion solutions were individually prepared from analytical grade Hg(II) chloride and Pb(II) nitrate. Stock solutions of 5 mmol/L (1000 mg/L Hg and 1040 mg/L Pb) were prepared in distilled water and slightly acidified with two to three drops of concentrated HNO3. 2.5. Batch adsorption studies 2.5.1. Sorbents’ dose The batch adsorption experiments were conducted in 100-mL Erlenmeyer flasks containing 50 mL of the tested metal solutions at an initial concentration of 0.5 mmol/L (100 mg/L Hg and 104 mg/L Pb) and the desired biochar and co-sorbents doses. The biochar doses ranged from 10 mg/L to 1000 mg/L. The co-sorbent was composed of series doses of B38 living biomass (a biomass concentration of 5×108 cells/mL, 0.1–5 mL) and 25 mg of biochar (CBC or PBC). Before the addition of the sorbents, the pH was adjusted to 5.0 with 1 M nitric acid and 1 M sodium hydroxide. The sorption experiments were conducted on a thermostatic shaker (Pui Ying HYG-A, China) operated at 20 °C and 150 rpm and lasted for 1440 min. After adsorption, the suspensions were centrifuged at 3000 rpm for 10 min. The supernatants were collected in the separate clean test tubes and analyzed for the residual metal concentrations in the aqueous phase. The controls were conducted under the same conditions using the metal ion solutions without sorbents. Each test was performed in triplicate.

2. Materials and methods 2.1. Microorganism, culture conditions and biosorbent preparation The bacterium used in this study was a mutant genotype (B38) from Bacillus subtilis, which was obtained by UV irradiation. This mutant species exhibits high tolerance to Cd, and it can tolerate 3 mmol/L (337 mg/L) Cd(II) in solution, whereas the original wild-type B. subtilis can only tolerate 0.25 mmol/L (28 mg/L) Cd(II) (Jiang et al., 2009). The gene sequences of the mutant species and the wild species were measured by Institute of Microbiology Chinese Academy of Sciences and are shown in Fig. A.1 in the Appendices A. After UV irradiation, a single base deletion occurred in B38 compared to the original wild type. In a former study, we checked sorption capacity of B38 on other heavy metals, and found that the maximum concentrations that B38 could tolerate were 4 mmol/L Cr, 0.5 mmol/L Hg, and 4.5 mmol/L Pb in solution (Wang et al., 2014). B38 has high affinities to Hg and Pb cations in solution (Wang and Sun, 2013). The bacterial suspension used as the biosorbent was with a biomass concentration of about 5 × 108 cells/ mL. Details of the culture conditions and the count method of the bacterial solution are the same as previously described by Wang and Sun (2013) and are presented in the Appendices A.

2.5.2. Sorption kinetics Batch adsorption experiments were conducted at the initial metal ion concentration of 0.5 mmol/L (100 mg/L Hg and 104 mg/L Pb) and with the certain sorbent dose (CBC or PBC, 500 mg/L; co-sorbent of 40 mg/L B38 biomass and 500 mg/L biochar). The samples were taken at different times (5, 10, 20, 30, 60, 120, 300, 600, 1200, 1440, 2400, and 2880 min) to obtain kinetic data. Other conditions employed were as described in Section 2.5.1.

2.2. Biochar and co-sorbent preparation

2.5.3. Sorption isotherms Batch adsorption experiments were conducted at the initial metal concentrations of 0.1–5 mmol/L (20–1000 mg/L Hg and 20.8–1040 mg/L Pb) and with the certain sorbent dose (CBC or PBC, 500 mg/L; co-sorbent of 40 mg/L B38 biomass and 500 mg/L biochar). The experiments lasted for 1440 min based on their kinetic data. Other conditions employed were as described in Section 2.5.1.

Corn straw stock was collected from a farmland in Jinnan District, Tianjin, China. Pig manure stock was collected from a hoggery in Jixian County, Tianjin, China. The corn straw and pig manure raw materials were air-dried, grounded to pass through a 2 mm (10 mesh) sieve, and then heated at 350 ◦C in a ceramic pot covered with a tight-fitting lid (where oxygen was soon exhausted) in a pre-heated muffle furnace for 2 h (Keiluweit et al., 2010). The produced biochars were grounded to pass through the 0.0380–0.075 mm (200–400 mesh) sieves and stored in a vacuum desiccator. The corn straw-derived biochar and pig manure-derived biochar are designated as CBC and PBC, respectively. The co-sorbent was composed of series doses of B38 living biomass (a biomass concentration of 5 × 108 cells/mL, 0.1–5 mL) and 500 mg/L of biochar (CBC or PBC), which were simultaneously added into solution immediately before the sorption experiment.

2.6. Competitive adsorption of cationic ions in binary system To investigate the possible competition between Hg(II) and Pb(II) ions, the adsorption experiment was conducted in a binary system. The sorbents (CBC or PBC, 500 mg/L; co-sorbents of 40 mg/L B38 biomass and 500 mg/L CBC or PBC) were brought into contact with a 50 mL 286

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(Park and Kim, 2004). The XPS results of the chemical compositions for the surface of the four sorbents surfaces are presented in Table 1. The oxygen content and the O1 s/C1 s composition ratio of CBC+B38 decreased due to their combination. In contrast, the O1 s peak intensity of PBC+B38 increased and the C1 s peak intensity decreased after combination. Meanwhile, the oxygen content and the O1 s/C1 s composition ratio of PBC+B38 increased due to their combination. These results clearly suggest the decrease of oxygen-containing functional groups on CBC surfaces after combination and the increase of that on PBC surfaces during the same process. To determine the forms of different complexes, the curve-fitting procedure on the O1 s peak of the samples were conducted. Fig. A.2b shows the fits to the XPS O1 s spectra of different samples. The O1 s subpeaks indicate the presence of several different species on each carbon: peak 1 (binding energy (BE) =531.6 eV) corresponds to C˭O groups (ketone, lactone, carbonyl), peak 2 (BE=533.3 eV) to -OH or CO-C groups, peak 3 (BE=534.0 eV) to O˭C-O groups (carbonyl), and peak 4 (BE=534.6 eV) to-C-O groups (Park and Kim, 2004). After the combination of biochar and B38, the peaks 3 and 4 disappeared. Thus, the combination mechanism of biochar and B38 is the heteropolar bonding between B38 and the carboxyl group on the biochar surface.

metal solution containing equimolar concentrations of each cation (0.5 mmol/L; 100 mg/L Hg and 104 mg/L Pb). 2.7. Analytical methods The concentrations of Pb(II) in the aqueous phase were determined using an atomic absorption spectrophotometer (AAS, Rayleigh WFX 210, China). The concentrations of Hg(II) in the aqueous phase were determined using an atomic fluorescence spectrometer (AFS, Jitian AFS 830, China). 2.8. Statistical analysis A one-way analysis of variance (ANOVA) at a P < 0.05 was used for determining the significance of the differences in the sorption capacities of different treatments. All the statistical tests were performed using OriginPro v8.5 software. 3. Results and discussion 3.1. Characterization of biochars

3.2. Metal sorption behavior

The bulk and surface chemical compositions of CBC and PBC are shown in Table 1. The ash content of pig manure-derived biochar (PBC) was much higher than that of corn straw-derived biochar (CBC). This result is consistent with the conclusions in the literature (Keiluweit et al., 2010). The high ash content in PBC was due to the high contents of mineral constituents in the feedstock (Brewer et al., 2011). In general, mineral impurities (ash) can serve as the additional adsorption sites for heavy metal adsorption (Xu et al., 2015). The atomic ratios of H/C and [(O+N)/C] are recognized as the indices for aromaticity and polarity, respectively (Chen et al., 2005). In the present study, the bulk H/C ratio of CBC was higher than that of PBC, while the trend for the bulk [(O+N)/C] ratio was opposite. Thus, CBC has lower bulk aromaticity and bulk polarity than PBC. However, the surface polarity [(O+N)/C] of CBC was greater than that of PBC, which was contrary to the trend in polarity for their bulk material. This indicated that CBC and PBC were of heterogeneous structures. The SEM photographs show substantial differences in structures between CBC and PBC (Fig. 1). During pyrolysis process, the unstable and volatile fractions in the raw materials disappeared. Thus, the carbon-skeleton structures of biochar were obviously distinct, and the porous structures were formed after carbonization. Moreover, CBC had a polyporous structure that is similar to sieve plates (Fig. 1a & b), while PBC had a polyporous structure with the lacunose surface (Fig. 1c & d). Thus, PBC could provide more adsorption sites for heavy metal binding and more space for microorganism growth. The bright dots in SEM photographs are the mineral fractions (ash) on the biochar surface (Fig. 1a & c). It could be seen that the ash content of PBC was higher than that of CBC, which is consistent with the results listed in Table 1. Fig. A.2a shows the XPS survey scan spectra of the four sorbents. The C1 s and O1 s peaks appear at 284.6 and 532.8 eV, respectively

3.2.1. Effect of the biochar dose on the adsorption of heavy metals The ratio of the adsorbed metal ion concentration at equilibrium to the initial metal ion concentration (the adsorption percentage (AP)), the amount of metal ion adsorbed per unit mass of adsorbent at time t (qt, mg/g), and at equilibrium (qe, mg/g), were calculated according to the following equations:

AP =

C0 − Ct × 100 C0

(1)

qt =

(C0 − Ct ) V m

(2)

qe =

(C0−Ce ) V m

(3)

where C0 (mg/L) is the initial metal ion concentration, Ct and Ce (mg/L) are the metal ion concentrations at time t and at equilibrium in the liquid phase, respectively, V (L) is the solution volume, and m (g) is the sorbent mass. The effect of biochar dosage on the equilibrium adsorption capacity of Hg(II) and Pb(II) ions was investigated (Fig. A.3). When the biochar was used as the sorbent alone, the AP increased from 4.3% to 98.9% as the biochar dosage increased from 10 to 1000 mg/L. At the highest dose of 1000 mg/L, the AP was over 95% for the two ions by CBC and PBC except in the case of Hg(II) binding by PBC, which was approximately 70%. Thus, a biochar dose (500 mg/L) with medium AP was chosen to further combine with B38 biomass to form co-sorbents. In the co-sorbent experiment, the AP first increased and then decreased with increasing B38 biomass loading dose (Fig. A.4). B38 has

Table 1 Elemental compositions and physic-chemical properties of CBC and PBC. Biochar

CBC CBC+B38 PBC PBC+B38 a b c

ash(%)

13.6 – 39.4 –

Bulk elemental composition (%)

Surface elemental composition (%)

C

H

N

O

C

N

O

P

S

63.6 – 39.6 –

3.6 – 2.2 –

0.9 – 5.1 –

18.5 – 13.8 –

64.1 74.7 69.5 65.8

9.0 1.4 4.5 8.9

25.7 23.1 22.1 22.6

0.7 0.4 3.0 1.9

0.6 0.4 0.9 0.9

pH, was measured using soil in 1:2.5 (w/v) 0.01 M CaCl2 solution. CEC, cation exchange capacity. OM, organic matter.

287

pHa

CECb(cmol/kg)

OMc(%)

7.5 – 8.1 –

108 – 64 –

94.8 – 52.7 –

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Fig. 1. SEM photography of CBC and PBC.

the high sorption capacities for Hg(II) and Pb(II) in solution. The theoretical Qmax based on Langmuir isotherm is 4.09 and 2.48 mmol/g for Hg(II) and Pb(II), respectively (Wang et al., 2014). The sorption capacities of the co-sorbents were enhanced with increasing B38 biomass loading doses as the B38 loading concentration increased up to 40 mL/ L. However, the AP remained steady or even decreased when the B38 loading dose was greater than 40 mL/L. This result is due to the reduction of the amount of available sorption sites on biochar surface by the occupation of B38. Therefore, a dose of 40 mL/L was chosen as the optimum loading dose of B38. 3.2.2. Sorption kinetics Sorption kinetics of the two heavy metal ions on the two biochars and the two co-sorbents are shown in Fig. 2. The adsorption process of the two metals occurred in two phases: a rapid adsorption in 60 min and then a relatively slow adsorption. The rapid metal binding at the beginning suggested that the metal ligands existed on the surface. Then, the metal ions were internalized by diffusion into the inner micropores of the biochar structure. After 300 min, the adsorption increased slowly and then achieved a steady state. The CBC exhibited a higher equilibrium adsorption capacity for Hg(II) than PBC did, while the PBC adsorbed Pb(II) ion better. As shown in Table 1, the CEC and OM contents of CBC were higher than PBC. In addition, the mean diameters of CBC and PBC particles in the bulk solution were investigated, and being 5.9 and 8.5 µm, respectively. These values can explain the higher affinity to Hg(II) of CBC than PBC. Cao et al. (2009) studied the Pb(II) adsorption mechanisms on dairy-manure-derived biochar by XRD and FTIR analyses. The sorption behavior was attributed to Pb(II) precipitation and surface sorption. Approximately 85% of the Pb ion was precipitated as β-Pb9(PO4)6 or Pb3(CO3)2(OH)2, and other 15% was attributed to surface sorption, probably via coordination of the Pb(II) d-electrons to C˭C (π electrons) and through -O-Pb bonds. Hence, P could be a key element

Fig. 2. The time-dependent adsorption of Hg and Pb ions by CBC, PBC or B38 living biomass combined with CBC or PBC.

for the adsorption of Pb(II). In the present study, the surface P composition of PBC was approximately 4-fold greater than CBC (Table 1). This can explain the greater Pb(II) sorption capacity of PBC than CBC. To investigate the sorption mechanisms, the characteristic constants of the adsorption kinetic were obtained using Lagergren's pseudo-firstorder equation (Eq. (4)) and pseudo-second-order equation (Eq. (5)):

1n (qe − qt ) = 1 nqe − kt t

(4)

⎛⎜ t ⎞⎟ = 1 + ⎜⎛ 1 ⎟⎞ t k2 qe2 ⎝ qt ⎠ ⎝ qe ⎠

(5)

where k1 (1/min) and k2 (g/(mg/min)) are the rate constant of the first288

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Table 2 Kinetics parameters obtained from pseudo-first-order (qe,exp : mg/g, k1: min−1) and pseudo-second-order (k2 :g/mg min, qe : mg/g) for Hg and Pb ions adsorption onto four different sorbents. Sorbents

CBC CBC+B38 PBC PBC+B38

qe,exp

Hg Pb Hg Pb Hg Pb Hg Pb

112.0 70.5 54.8 87.2 94.4 103.9 72.7 116.0

Pseudo-first-order kinetics

Pseudo-second-order kinetics

k1

qe1, cal

r2

k2

qe2, cal

r2

0.040 0.162 0.057 0.052 0.057 0.045 0.222 0.044

104.2 60.2 50.7 68.9 95.0 98.4 53.9 96.4

0.9333 0.8284 0.7807 0.6112 0.9913 0.9037 0.6610 0.6896

0.001 0.004 0.002 0.001 0.001 0.001 0.006 0.001

89.5 63.3 53.2 73.9 99.0 102.5 56.0 101.3

0.8373 0.9123 0.8704 0.7200 0.9673 0.9681 0.6982 0.8065

Table 3. Higher regression correlation coefficients (r2 > 0.90) were obtained for the Freundlich isotherms than for the Langmuir and Langmuir-Langmuir isotherms in the adsorption process by the two biochars. All non-linearity values (n) of biochar sorption isotherms were less than 1.0. Hence, the adsorption of Hg(II) and Pb(II) by biochar was a multi-layer sorption process with the heterogeneous energetic distribution of active sites (Kim et al., 2013; Komnitsas et al., 2015). The n values of PBC were lower than CBC, which implied that PBC was more heterogeneous than CBC. The maximum adsorption capacities (Qmax) of Hg(II) based on the Langmuir isotherms for CBC and PBC were comparable (151.5 mg/g and 158.7 mg/g for CBC and PBC, respectively). However, the Qmax of Pb(II) for PBC was much higher than CBC (113.6 mg/g and 151.5 mg/g for CBC and PBC, respectively) (Table 2). Moreover, the Qmax values of CBC and PBC in this study are larger than other adsorbents in the literatures, such as activated carbon (Ranganathan, 2003; Wang et al., 2010) and carbon nanotubes (Lin et al., 2012; Hadavifar et al., 2014). Thus, the results clearly demonstrated that the biochars have the potential to be an effective adsorbent for Hg(II) and Pb(II) removal from solution. The Langmuir-Langmuir isotherm is most suitable (r2 > 0.97) for describing the adsorption equilibrium of heavy metals by the two cosorbents in the studied concentration range (0.1–5 mmol/L). Dual models have been successfully used to describe the metal sorption isotherms involving multiple sorption processes, e.g., surface adsorption, precipitation and complexation (Das et al., 2007). It was revealed that the surface characteristics of biochar changed after combining with B38. Moreover, there were two different primary adsorption mechanisms involved in the adsorption process of Hg(II) and Pb(II) by the cosorbents. The Qmax ,2 is much higher than Qmax ,1 in the LangmuirLangmuir model (Table 3), indicating that precipitation was mainly responsible for Hg(II) and Pb(II) adsorption by the co-sorbents (76–91%). The remaining 9–24% of the sorption (Qmax ,1) was attributed to surface sorption. The precipitation accounted for over 90% of the Hg(II) adsorption. Besides, the Qmax values of the co-sorbents were much higher than CBC and PBC, and other different types of biochars reported in the literatures (Cao et al., 2009; Kong et al., 2011; Xu et al., 2015).

order equation and the second order equation, respectively (Özdemir et al., 2009). The k1, r2 and qe values are given in Table 2. These constants show that the sorption of Hg(II) by the two biochars was well fit to the pseudo-first-order kinetics model (r2 values of CBC and PBC were 0.9333 and 0.9913, respectively), but the sorption process of Pb was better fit by the pseudo-second-order kinetics model (r2 values of CBC and PBC were 0.9123 and 0.9681, respectively). Furthermore, the theoretical maximum uptake values (qe, cal) agreed very well with the experimental data (qe, exp) for the pseudo-first-order kinetics model. The pseudo-first-order kinetics model is a chemical kinetics model that is based on the relationship between the reactant concentration and the reaction rate. Previous studies have shown that the application of the pseudo-first-order kinetics model was successful when the sorption process was rapid (Kumar and Sivanesan, 2006; Greluk and Hubicki, 2010). The pseudo-second-order kinetics model describes a complex sorption process with multiple steps. Many adsorption reactions occur through a multi-step mechanism comprising: (i) external surface diffusion, (ii) intraparticle diffusion, and (iii) the interaction between the adsorbate and the active sites. Because the first step is excluded by shaking the solution, the rate-determining step is one of the other two steps (Ji et al., 2011). Thus, the process of Hg(II) ion adsorbed on biochar was a rapid process. In contrast, the mechanisms of Pb(II) sorption by biochar were complex and included some rate-limiting steps. The adsorption kinetics of the two metals on the two co-sorbents were different from those of the two biochars. Rapid adsorption occurred in the initial 60 min, and the AP reached approximately 50%. Then, the adsorption rate slowed down, but the AP continued to increase. The adsorption process achieved steady-state after 1200 min. This phenomenon might be because after a period of adaptation to the new culture conditions, the B38 cells from the late log-phase of culture growth achieved the maximum activation. The sorption capacity of Pb (II) was enhanced on the co-sorbents as compared to the biochar alone, but that of Hg(II) was weakened (Fig. 2). It was deduced that the B38 cells imported new sorption sites for Pb(II), but blocked some sorption sites existed on biochar for Hg(II). The sorption processes of Hg(II) and Pb(II) by the two co-sorbents were fitted well to the pseudo-secondorder kinetics model. These results suggested that several processes were involved in the sorption of the co-sorbents, including some ratelimiting steps, such as the bacterial growth.

3.2.4. Binary system sorption To illustrate the sorption behavior in the bi-solute system, the q′e/qe ratios were calculated, where qe and q′e are the equilibrium sorption capacities in single and binary sorption systems, respectively. In general, three possible types of interaction are exhibited: q′e/qe > 1 indicates synergism; q′e/qe < 1 indicates antagonism and q′e/qe =1 indicates non-interaction. The q′e/qe ratios for the sorption of one metal in the presence of another metal are shown in Table A1. In the binary system, the adsorption capacity of Pb by the four sorbents was higher than that of Hg. The q′e/qe ratios for the sorption of Hg(II) in the presence of Pb(II) by CBC and PBC were 0.57 and 0.66, respectively, whereas the q′e/qe ratios

3.2.3. Sorption isotherms Generally, the adsorption isotherms provide vital information to optimize the use of adsorbents. The Langmuir, dual Langmuir-Langmuir and Freundlich adsorption isotherms were used in this study to fit the adsorption data. Details of the three isotherm models are described in Appendices A. The sorption isotherms of Hg(II) and Pb(II) onto the four sorbents are presented in Fig. A.5, and the fitting parameters are listed in 289

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Table 3 Adsorption isotherm parameters for Hg and Pb by the four sorbents in solution. Metal

Hg

Pb

Sorbent

CBC CBC+B38 PBC PBC+B38 CBC CBC+B38 PBC PBC+B38

Freundlich isotherm Qe = KF ⋅Cen

Langmuir-Langmuir isotherm

Qe =

Qmax ,1 ⋅ KL,1 ⋅ Ce 1 + KL,1 ⋅ Ce

+

Langmuir isotherm

Qmax ,2 ⋅ KL,2 ⋅ Ce

Qe =

1 + KL,2 ⋅ Ce

Qmax ⋅ KL ⋅ Ce 1 + KL ⋅ Ce

KF

n

r2

Qmax,1 (mg/g)

KL,1 (L/mg)

Qmax,2 (mg/g)

KL,2 (L/mg)

r2

Qmax (mg/g)

KL (L/mg)

r2

24.95 17.19 27.50 28.00 38.99 53.72 63.74 64.34

0.379 0.364 0.366 0.333 0.233 0.213 0.205 0.274

0.9415 0.9588 0.9682 0.9392 0.9071 0.9364 0.9274 0.9738

357.0 39.8 522.4 54.6 272.1 86.2 388.4 54.7

0.004 0.258 0.003 0.728 0.007 0.222 0.006 0.725

289.4 394.2 17.4 536.5 219.5 279.3 13.8 274.0

1.2*10−3 1.1*10−3 −7.0*10−5 1.2*10−3 −1.1*10−4 2.1*10−3 −6.0*10−5 5.9*10−3

0.9082 0.9750 0.8945 0.9925 0.8537 0.9847 0.8472 0.9741

151.5 97.1 158.7 126.6 113.6 126.6 151.5 166.7

0.20 0.17 0.16 0.31 1.04 11.29 5.50 4.62

0.8974 0.8579 0.9035 0.9027 0.7197 0.8454 0.8611 0.8888

for the sorption of Pb(II) in the presence of Hg(II) by CBC and PBC were 0.71 and 0.73, respectively. The ratios of Hg and Pb in the binary system with the two biochars were all < 1, indicating that the adsorption of each metal was depressed by the presence of another metal ion; therefore, the effect of the mixtures appeared to be antagonistic. Moreover, the suppression effect of Pb(II) on Hg(II) was much greater than that of Hg(II) on Pb(II). These findings could be explained by the physic-chemical properties of the ions. The preference of the biochar for Pb(II) over Hg(II) could be because Pb(II) is paramagnetic and has two possible coordination modes (2 or 4), besides, Pb(II) is more electronegative and has a larger atomic weight than Hg(II). Moreover, Cao et al. (2009) found that the main sorption mechanism of Pb by dairymanure-derived biochar was the precipitation with PO43- ion in biochar. Both CBC and PBC contained P in their surface elemental compositions (Table 1). Thus, the adsorption of Pb(II) by biochar gained the competitive advantage over Hg(II) in the binary system. The overall adsorption capacities of biochar for Hg-Pb (CBC, 113.8 mg/g; PBC, 140.8 mg/g) were higher than the adsorption capacities of the two ions alone, suggesting that the adsorption sites of Pb(II) may partially overlap those of Hg(II) in biochar. The q′e/qe ratios in the Hg-Pb binary systems of both co-sorbents were all ≈1, indicating that their co-existence has almost no effect on the adsorption of each metal ion in the mixture (Table A1). The overall adsorption capacity for Hg-Pb (CBC+B38, 136.7 mg/g; PBC+B38, 181.3 mg/g) was greater than the single-component values for Hg(II) (CBC+B38, 57.4 mg/g; PBC+B38, 72.7 mg/g) or Pb(II) (CBC+B38, 87.2 mg/g; PBC+B38, 116.0 mg/g). It was approximately equal to the sum of the single-component values of Hg(II) and Pb(II) (CBC+B38, 144.6 mg/g; PBC+B38, 188.7 mg/g). Thus, the adsorption capability of the co-sorbents was much higher than biochars without B38. This capability is an advantage of the co-sorbent; they are more effective in removing heavy metals from a mixture. Hg(II) has an affinity for organic sulfur groups called thiols, with a decreasing affinity for other groups in the following sequence: sulfur, amides, amines, carbon, and phosphate (Gavis and Ferguson, 1972). Sulfur is present in all proteins, which makes it universally available for binding with Hg(II). However, Pb(II) has an affinity for the phosphate and carbonate in the biochar surface (Cao et al., 2009), and binds to the predominant phosphoryl and minor carboxyl groups in the cell walls (Sarret et al., 1998). In our previous study, it was found that the sorption sites of Hg(II) and Pb(II) on the surface of B38 cells were different (Wang and Sun, 2013). This difference implied that B38 imported some specific sorption sites for Hg(II) and Pb(II) onto CBC and PBC during the combination process. Hence, the sorption sites of Hg(II) and Pb(II) may partially overlap on the biochar surface, but are different on the co-sorbent surface. Fig. 3. FT-IR spectrogram of the four sorbents ( a: CBC and PBC before and after combination with B38 mutant; b: CBC and PBC before and after adsorption with Hg(II) or Pb (II); c: CBC+B38 and PBC+B38 before and after adsorption with Hg(II) or Pb(II)).

3.2.5. FTIR spectral analysis The FTIR spectra of the four sorbents with and without metal ions were investigated (Fig. 3). The FTIR fingerprints appear to be quite 290

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mutant species; (2) XPS spectra and high-resolution O1 s XPS spectra of 4 different sorbents; (3) Effect of the sorbent dose on the equilibrium biosorption capacity of Hg and Pb ions by CBC and PBC; (4) Effect of the B38 sorbent dose on the equilibrium biosorption capacity of Hg and Pb ions by B38 living biomass combined with CBC or PBC; (5) Freundlich and Langmuir-Langmuir adsorption isotherms of Hg(II) and Pb(II) adsorption on 4 different sorbents. 2. Details of the culture and count method of the bacterial solution. 3. Details of the Langmuir, dual Langmuir-Langmuir and Freundlich adsorption isotherm models. 4. One table showing the equilibrium adsorption in single and binary systems.

different for biochar with and without B38 (Fig. 3a). After the combination with B38, the intensity and the amplitude of most peaks from the CBC decreased, and the peaks became broader. In contrast, the intensity and the amplitude of most of the PBC peaks increased, and the peaks became narrower after combining with B38. These changes suggested that the amount of functional groups on the CBC surface decreased after the combination with B38, while that of PBC increased. This difference in functional groups could explain the change in the sorption capacity of the two co-sorbents. After the adsorption of metal ions on the two biochars, the intensity of the peak at 3700-3000 1 /cm obviously decreased, and the peak became broader (Fig. 3b). These changes suggested that the hydroxyl functional groups were involved in the interaction with heavy metals ions via the ion exchange reaction (Das et al., 2009). The crest of the peak at 1260 1 /cm became broader, and the band was redshifted, which revealed the interaction of heavy metals with the C˭O bond of aromatic and the –OH bond in phenol (Chun et al., 2004). The peak at 1160-1030 1 /cm was redshifted by approximately 1/20 cm, and the intensity of the peak obviously decreased. This suggests the interaction of the heavy metals with the groups from the cellulose in CBC (Bustin and Guo, 1999). The peak at 900-700 1 /cm is due to the vibration of the γ-CH of furans and the β-rings of pyridines (Das et al., 2009). After the adsorption, the intensity of the peak at 900-700 1 /cm decreased sharply. Thus, it could be deduced that the cation-π interaction was involved in the adsorption of heavy metal ions by biochars. As for the co-sorbents, after the adsorption of Pb(II), the intensity and the amplitude of most of the peaks in CBC+B38 decreased (Fig. 3c). This reduction suggests that the hydroxyl, carbonyl, carboxyl, γ-CH of furans, and β-rings of pyridines were involved in the interaction with Pb(II) ion by the ion exchange reaction and cation-π interactions. These results are in good accordance with the results obtained in the batch adsorption experiments in that combining CBC with B38 enhanced the adsorption of Pb(II) compared to biochar alone. However, after the adsorption of Hg(II), there were no obvious changes in the intensity, amplitude or peak location. Thus, there were almost no changes in the functional groups in CBC+B38 surface. CBC+B38 had relatively low affinity to Hg(II). Thus, the main adsorption mechanisms were the ion exchange, electrostatic interaction and precipitation. After the adsorption of Hg(II) and Pb(II), the intensity and the amplitude of most peaks of PBC+B38 decreased. This observation suggests that the hydroxyl, carbonyl, carboxyl, γ-CH of furans, and β-rings of pyridines were involved in the interaction by the ion exchange reaction and cation-π interaction (Das et al., 2009; Xu et al., 2015).

Acknowledgements This work was financially supported by Natural Science Foundation of Tianjin [15JCQNJC15200] and Key Technologies R & D Program of Tianjin [16YFZCSF00410]. Appendix A. Supporting information Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.ecoenv.2017.10.039. References Al-Hakawati, M., Banks, C., 2000. Copper removal by polymer immobilised Rhizopus oryzae. Water Sci. Technol. 42, 345–352. Arıca, M.Y., Arpa, Ç., Ergene, A., Bayramoğlu, G., Genç, Ö., 2003. Ca-alginate as a support for Pb (II) and Zn (II) biosorption with immobilized Phanerochaete chrysosporium. Carbohyd. Polym. 52, 167–174. http://dx.doi.org/10.1016/S0144-8617(02) 00307-7. Arshadi, M., 2015. Manganese chloride nanoparticles: a practical adsorbent for the sequestration of Hg(II) ions from aqueous solution. Chem. Eng. J. 259, 170–182. http:// dx.doi.org/10.1016/j.cej.2014.07.111. Azargohar, R., Dalai, A., 2006. Biochar as a precursor of activated carbon. Appl. Biochem. Biotech. 131, 762–773. http://dx.doi.org/10.1385/ABAB:131:1:762. Brewer, C.E., Unger, R., Schmidt-Rohr, K.R., Brown, C., 2011. Criteria to select biochars for field studies based on biochar chemical properties. Bioenerg. Res. 4, 312–323. http://dx.doi.org/10.1007/s12155-011-9133-7. Bustin, R., Guo, Y., 1999. Abrupt changes (jumps) in reflectance values and chemical compositions of artificial charcoals and inertinite in coals. Int. J. Coal Geol. 38, 237–260. http://dx.doi.org/10.1016/S0166-5162(98)00025-1. Cao, X., Ma, L., Gao, B., Harris, W., 2009. Dairy-manure derived biochar effectively sorbs lead and atrazine. Environ. Sci. Technol. 43, 3285–3291. http://dx.doi.org/10.1021/ es803092k. Cao, X., Ro, K.S., Chappell, M., Li, Y., Mao, J., 2010. Chemical structures of swine-manure chars produced under different carbonization conditions investigated by advanced solid-state 13C nuclear magnetic resonance (NMR) spectroscopy. Energ. Fuel. 25, 388–397. http://dx.doi.org/10.1021/ef101342v. Chen, B.L., Johnson, E.J., Chefetz, B., 2005. Sorption of polar and nonpolar aromatic organic contaminants by plant cuticular materials: the role of polarity and accessibility. Environ. Sci. Technol. 39, 6138–6146. http://dx.doi.org/10.1021/es050622q. Chen, T., Zhang, Y., Wang, H., 2014. Influence of pyrolysis temperature on characteristics and heavy metal adsorptive performance of biochar derived from municipal sewage sludge. Bioresour. Technol. 164, 47–54. http://dx.doi.org/10.1016/j.biortech.2014. 04.048. Chun, Y., Sheng, G., Chiou, C.T., Xing, B., 2004. Compositions and sorptive properties of crop residue-derived chars. Environ. Sci. Technol. 38, 4649–4655. http://dx.doi.org/ 10.1021/es035034w. Coral, M.N.U., Korkmaz, H., Arikan, B., Coral, G., 2005. Plasmid mediated heavy metal resistances in Enterobacter spp. isolated from Sofulu landfill, in Adana, Turkey. Ann. Microbiol. 55, 175–179. Das, D.D., Schnitzer, M.I., Monreal, C.M., Mayer, P., 2009. Chemical composition of acid–base fractions separated from biooil derived by fast pyrolysis of chicken manure. Bioresour. Technol. 100, 6524–6532. http://dx.doi.org/10.1016/j.biortech.2009.06. 104. Das, S.K., Das, A.R., Guha, A.K., 2007. A study on the adsorption mechanism of mercury on Aspergillus versicolor biomass. Environ. Sci. Technol. 41, 8281–8287. http://dx. doi.org/10.1021/es070814g. Gavis, J., Ferguson, J.F., 1972. The cycling of mercury through the environment. Water. Res. 6, 989–1008. http://dx.doi.org/10.1016/0043-1354(72)90053-X. Glaser, B., Birk, J.J., 2012. State of the scientific knowledge on properties and genesis of Anthropogenic Dark Earths in Central Amazonia (terra preta de Índio). Geochim. Cosmochim. Acta 82, 39–51. http://dx.doi.org/10.1016/j.gca.2010.11.029. Greluk, M., Hubicki, Z., 2010. Kinetics, isotherm and thermodynamic studies of Reactive Black 5 removal by acid acrylic resins. Chem. Eng. J. 162, 919–926. http://dx.doi. org/10.1016/j.cej.2010.06.043.

4. Conclusions Biochar derived from corn straw and pig manure, showed great sorption capacity for Hg(II) and Pb(II) in solution, with the theoretical maximum adsorption capacity reaching approximately 150 mg/g. Meanwhile, the two biochars can act as good carrier for microorganisms to form co-sorbents. After combination with B38, the sorption capacity of the two co-sorbents for Pb(II) was enhanced as compared to the biochar alone, but which for Hg(II) was weakened. Electrostatic interaction and precipitation were the major mechanisms in the adsorption of Hg(II) on the co-sorbent. In contrast, cation-π interactions and precipitation were involved in the adsorption of Pb(II). The sorption sites of Hg(II) and Pb(II) partially overlapped on the biochar surface, but were different in the co-sorbents samples. The co-sorbents showed an advantage over the biochar alone in the adsorption of Hg(II) and Pb(II) in a binary system. The co-sorbents of biochar and microorganisms provides a promising approach immobilize multiple heavy metal contaminants from waste wasters and contaminated soil. Supplementary information 1. Five figures showing (1) DNA fragments of the wild species and 291

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