Characterization and Pb(II) removal potential of corn straw- and municipal sludge-derived biochars

Corn straw- and municipal sludge-derived biochars (CS-BC and MS-BC, respectively) were used to remove Pb(II) from aqueous solutions. Despite being pyrolysed at the same temperature (723 K), MS-BC showed higher porosity and hydrophobicity than CS-BC. The optimum biochar loading and pH values allowing efficient Pb(II) removal (greater than 80%) were 0.2 g l−1 and 7.0, respectively. The presence of PO43− (greater than 0.01 mol l−1) significantly affected the adsorptive performance of Pb(II) on the biochar samples. The adsorption data fitted well to a pseudo-second-order kinetic model and a Langmuir model, and the maximum Pb(II) adsorption capacities were 352 and 387 mg g−1 for CS-BC and MS-BC, respectively. The main mechanisms involved in the adsorption of Pb(II) on biochar were electrostatic attraction and surface complexation. When comparing both biochars, CS-BC showed better cost-effectiveness for the removal of Pb(II) from aqueous solutions.

throughout this study. A Pb(II) stock solution (1000 mg l −1 ) was prepared by dissolving Pb(NO 3 ) 2 in ultrapure water. The pH of the experimental solutions was adjusted using 0.1 mol l −1 NaOH or HNO 3 solutions. NaNO 3 , Na 2 CO 3 and Na 3 PO 4 stock solutions (1 mol l −1 ) were prepared by dissolving the corresponding amount of the chemicals in ultrapure water.

Characterization of biochars
The as-prepared CS-BC and MS-BC samples were characterized by elemental analysis (Vario EL, German Elementar Co., Germany), scanning electron microscopy (SEM, S250MK3, Cambridge UK Co., UK), Fourier transform infrared spectroscopy (FTIR, Germany BRUKER Spectrometer Co., Germany) and X-ray diffraction (XRD, X'Pert PRO MPD, Holland Research Co., The Netherlands). Elemental analysis was used to determine the C, H, N, O and S contents of the two different biochars. The Brunauer-Emmett-Teller (BET) surface areas and the micropore volumes (MV) were determined from N 2 adsorption isotherm data obtained at 77 K on an accelerated surface area and porosimetry system (ASAP 2020, Micromeritics, USA). SEM was used to determine the structure of the two different biochars. XRD analysis of two different biochars was carried out on a diffractometer provided with Cu Kα radiation (Kα = 1.54 nm) at a voltage of 40 kV and a current of 40 mA [24]. FTIR analysis (400-4000 cm −1 ) was used to determine the surface functional groups of CS-BC and MS-BC before and after adsorption. The ash content of the biochars was determined by combusting the samples in a muffle furnace at 873 K for 4 h and subsequent cooling in a desiccator until constant weight. Moreover, the zeta potential was determined at different pH values using a potential analyser (Zetasizer Nano, UK), while the pH of the biochar samples was measured by adding biochar to ultrapure water at a mass : water ratio of 1 : 20.

Batch adsorption experiments
Batch-mode adsorption studies were conducted to investigate the effects of the biochar loading, pH and coexisting anions on the Pb(II) adsorption process. CS-BC and MS-BC were added to 10 ml polyethylene (PE) centrifuge tubes containing a Pb(II) solution with an initial concentration of 40 mg l −1 at varying solid mass : liquid volume ratios (i.e. 0.0, 0.02, 0.04, 0.08, 0.1, 0.2, 0.4, 0.6, 0.8 and 1.0 g l −1 ) and the resulting solutions were shaken at 150 r.p.m. and 298 K for 8 h in a vertical temperature oscillation incubator (ZQPL-200,Tianjin Lai Bo Terry instrument Equipment Co., Tianjin, China). The suspensions were filtered with a 0.45 µm polysulfone filter membrane. The concentration of Pb(II) remaining in the supernatant solution was measured by inductively coupled plasma mass spectrometry (ICP-MS, Agilent 7500, USA) with a good method recovery (98 ± 8.7%) and a low detected limit (2 ng l −1 ). The optimal loading was selected for further experiments. To investigate the influence of the pH, the initial pH of the Pb(II) solutions was previously varied from 2 to 11 by using 0.1 mol l −1 NaOH or HNO 3 solutions before performing the adsorption experiments as indicated above. The effect of coexisting anions (i.e. NO − 3 , CO 2− 3 and PO 3− 4 ) on the adsorption of aqueous solutions of Pb(II) over biochar was studied by adding NaNO 3 , Na 2 CO 3 and Na 3 PO 4 at varying concentrations (0.001-0.1 mol l −1 ) in the initial Pb(II) solution. Every batch experiment was set with three parallel samples, and a blank solution experiment was conducted under the same test procedure. The removal efficiency and the adsorbed amount of Pb(II) ion at equilibrium were calculated using equations (2.1) and (2.2), respectively [25]: and where C 0 is the initial concentration of Pb(II) (mg l −1 ); C e represents the equilibrium concentration of Pb(II) (mg l −1 ); q e is the amount of adsorbed Pb(II) (mg g −1 ); V is the volume of Pb(II) solution (l) and m is the weight of CS-BC or MS-BC (g).

Adsorption kinetics
The kinetics for the adsorption of metal ions can be used to identify the main adsorption mechanism. Based on the above experiments, the optimal biochar loading, pH and anions concentration were selected for performing the adsorption kinetics, while the initial concentration of Pb(II) was fixed to 40  used to describe the adsorption process of Pb(II) ion adsorption on CS-BC and MS-BC. The pseudo-firstorder and pseudo-second-order models can be expressed by the following equations [26]: and (pseudo-second-order), (2.4) where, q t and q e (mg g −1 ) are the amounts of metal ions adsorbed at contact time t (min) and at equilibrium, respectively; K 1 is the rate constant of the pseudo-first-order adsorption model (min −1 ); and K 2 is the rate constant of the pseudo-second-order adsorption model (g mg −1 min −1 ).

Adsorption isotherm
For the adsorption isotherm study, Pb(II) aqueous solutions with varying initial concentrations (i.e. 20, 40, 60, 80, 100, 120, 140, 160, 180 and 200 mg l −1 ) were used at optimal pH, biochar loading, coexisting anion concentration and equilibrium time conditions previously determined. Furthermore, in order to study the effect of the temperature on the adsorption of Pb(II), adsorption experiments were conducted at 298, 313 and 328 K. Freundlich and Langmuir adsorption isotherm models were employed to fit the Pb(II) adsorption data on biochar. Langmuir and Freundlich isotherm models are, respectively, described as follows [27]: and where Q max is the maximum adsorption capacity of the biochar sample (mg g −1 ); q e is the equilibrium adsorption capacity of the biochar sample (mg g −1 ); b is the Langmuir adsorption characteristic constant (L mg −1 ); K F and n represent the Freundlich empirical constants (l g −1 ) and C e is the adsorption equilibrium concentration (mg l −1 ).

Characterization of the biochar
The physico-chemical properties of CS-BC and MS-BC (i.e. productivity, ash content, pH and elemental composition) are listed in table 1. CS-BC showed a yield (44.4%) significantly higher than that of MS-BC (26.3%). The ash content of CS-BC (9.24%) was notably lower than that of MS-BC (48.7%) mainly because of the decomposition of volatile substances (CO 2 ) and accumulation of minerals (KHCO 3 ) at high content in the latter sample [9]. MS-BC showed higher BET surface area and MV as compared to CS-BC (table 1). There are the higher cellulose and lignin contents of corn straw than those of municipal sludge. High lignin-content biomass is not easily decomposed at low pyrolysis temperatures (less than 798 K), thereby resulting in incompletely developed pore structures. CS-BC and MS-BC were both weakly alkaline (average pH values of 8.18 and 7.36, respectively). CS-BC and MS-BC also showed low pH pzc values (less than 5.0), thereby revealing high acidic characteristics and thus strong buffer capacity under basic environments [28]. The carbon content of CS-BC was significantly higher than that of MS-BC, which is consistent with the higher carbon content reported for other agricultural sourcederived biochars [3,29]. The O/C and H/C ratios can be uses as an indication of the hydrophilicity and carbonization degree of biochar, respectively [30]. The higher value of O/C and H/C further confirmed that these biochars were incompletely decomposed under this pyrolysis temperature. Compared with CS-BC, MS-BC showed higher hydrophobicity and carbonization degree, thereby revealing a higher number of surface adsorption sites available for this biochar [3]. SEM micrographs and XRD patterns of CS-BC and MS-BC are shown in figure 1. As shown in figure 1a,c, CS-BC showed a coarse fibre surface structure (4.36 m 2 g −1 ), thereby revealing a significantly higher fraction of hemicellulose and cellulose as compared to MS-BC that showed a more smooth and slightly crumby structure with higher surface area (10.1 m 2 g −1 ). Compared with CS-BC, some bright zones or points were clearly observed on the surface of MS-BC revealing the presence of Fe, Al and Si as determined by XRD analysis (figure 1d). These results were consistent with the findings of Singh      et al. [31], who reported the presence of quartz, kaolinite, haematite and sylvite in cow manure-and poultry litter-derived biochars pyrolysed at 823 K. These minerals can co-precipitate along with heavy metal ions, thereby enhancing metal adsorption [3]. In the case of CS-BC, higher amounts of K + and Ca 2+ were observed (figure 1b), thereby allowing Pb(II) ions to be adsorbed on the biochar via electrostatic cation exchange with these ions or metal exchange reactions (i.e. surface complexes or precipitation) [32].

Effect of the biochar loading
The biochar loading is an important factor during the adsorption process. As shown in figure 2a, both biochars showed similar change trends (i.e. the Pb(II) adsorption capacity increased with the biochar loading). In the case of MS-BC, the Pb(II) adsorption capacity rapidly increased up to 123 ± 0.21 mg g −1 with the dosage increasing from 0.0 to 0.1 g l −1 and slowly decreased thereafter (up to a biochar loading of 1.0 g l −1 ). In the case of CS-BC, the same trend was observed for biochar loadings of 0.0-0.2 g l −1 (largest Pb(II) adsorption capacity value of 84.8 ± 0.30 mg g −1 ). Although larger biochar loadings can provide more active sites for adsorption, the adsorption capacity of the biochar was observed to decrease at loadings above a certain value. At these conditions, biochar aggregation occurs thereby reducing the number of binding sites while also favouring electrostatic repulsion between the biochar and the metal ions [24]. The Pb(II) adsorption capacity over MS-BC was significantly higher than that of CS-BC because of the higher porosity and the presence of surface functional groups to a significantly larger extent in the former. CS-BC and MS-BC both showed their maximum Pb(II) adsorption capacity at a biochar loading of 0.2 g l −1 . Thus, this biochar loading was selected herein as the optimum as it allows minimum utilization of the biochar samples, full use of their adsorption capability, and reasonable comparison between the adsorption behaviour of MS-BC and CS-BC.  figure 2b, the Pb(II) adsorption on MS-BC and CS-BC was dependent on the initial pH value of the solution, with 7 being the optimal pH conditions. The Pb(II) adsorption capacity on MS-BC increased up to 172 ± 0.3 mg g −1 with the pH varying in the range of 2-7 and slightly decreased thereafter. In the case of CS-BC, no adsorption of Pb(II) was observed at low pH values (2)(3)(4). The amount of Pb(II) adsorbed increased up to 157 ± 0.4 mg g −1 with the pH varying in the 4-7 range, remained constant thereafter (pH = 7-10), and decreased at pH values higher than 10. At low pH values, MS-BC (pH pzc = 2.08) and CS-BC (pH pzc = 4.05) were positively charged on the surface, and high electrostatic repelling forces inhibited the contact of Pb(II) ions and the biochar surface. At pH values higher than pH pzc , the electrostatic attraction between the biochar surface and the metal ions is enhanced because the biochar is negatively charged on its surface. Moreover, the concentration of H + ions in solution and thus their competition with Pb(II) for surface adsorption sites decreases with the pH, thereby positively affecting the adsorption capacity of Pb(II). However, at higher pH values (pH > 6), Pb(II) can precipitate upon reaction with the OH − ions, thereby reducing its mobility and leading to a lower adsorption capacity of MS-BC. The Pb(II) adsorption capacity of CS-BC remained constant in the pH range of 7-10, and this may be attributed to the protonation-deprotonation of carboxyl and hydroxyl groups on the biochar in the pH region [28]. The electrostatic interactions probably play an important role in controlling the removal of Pb(II) ions over biochar from aqueous solutions at different pH values, especially in the case of MS-BC.

Effect of coexisting anions
The

Effect of the surface functional groups
The FTIR spectra of the two biochars were measured before and after Pb(II) adsorption (figure 3). As shown in figure 3, before Pb(II) adsorption both biochars exhibited absorption bands at 3407 cm −1 (−OH functional groups), 2919, 2852 and 796 cm −1 (−CH functional groups), 1611 cm −1 (C=C functional groups) and 1375 cm −1 (−COOH functional groups) [1,33]. Moreover, stretching P=O bands at 1242 cm −1 (CS-BC) and C=O bands at 1040 cm −1 (MS-BC) were observed. After Pb(II) adsorption, these bands corresponding to −OH, −CH and −COOH groups were notably displaced for both biochars, thereby indicating that the Pb(II) adsorption mechanisms were surface complexation (with carboxyl and hydroxyl functional groups) and coordination with the π electrons of aromatic −CH groups [34,35]. In the case of MS-BC, other functional groups such as C=C and C=O showed noticeable vibrations, thereby suggesting that a significantly higher number of functional groups were involved in the Pb(II) removal via surface complexation and coordination. Similar results were reported for Pb(II) and Cd(II) adsorption on a sludge-derived biochar [14,24], thereby implying that the functional groups play an important role during the adsorption of heavy metals in aqueous solutions.  equilibrium. Pseudo-first-order and pseudo-second-order kinetic models can be used to describe the adsorption of heavy metals [36]. The K 1 , K 2 , q e and R 2 (correlation coefficient) values for these models are shown in table 2. As can be seen in table 2, unlike the pseudo-second-order kinetic model (R 2 > 0.98), the correlation coefficients for the pseudo-first-order kinetic model were low (R 2 < 0.80). Thus, a pseudosecond-order model was more suitable to describe the adsorption of Pb(II) on biochars, thereby revealing that chemical interactions between Pb(II) and the surface adsorption sites occurred during the adsorption process [35].   As confirmed by previous reports, the adsorption of Pb(II) on agricultural biochars is an endothermic process [33] and, therefore, the Pb(II) adsorption capacity increased with temperature.

Adsorption isotherms and the effect of temperature
To further evaluate the interactions between Pb(II) and biochars while varying the temperature, thermodynamic parameters such as the Gibbs free energy change ( G 0 ), the standard enthalpy ( H 0 ) and the standard entropy ( S 0 ) were calculated using the following equations: where R is the gas constant (8.314 J K −1 mol −1 ), T is the absolute temperature in Kelvin and K is an equilibrium constant obtained by multiplying the Langmuir constants Q max and b. The calculated thermodynamic parameters are listed in table 3. The negative values of G 0 suggested that the adsorption of Pb(II) on CS-BC and MS-BC was a thermodynamically favourable and spontaneous process. The positive value of S 0 revealed an increase in the disorder of the solid solution system [37]. The positive value of H 0 confirmed that the adsorption of Pb(II) on both biochars was an endothermic process. The high values obtained might account for the energy required to destroy the hydration sheath of Pb(II) with molecular water and to form some chemical bonds between Pb(II) and the functional groups in biochar during the adsorption process [37].  biochars tested herein (CS-BC: 288 mg g −1 and MS-BC: 289 mg g −1 ) was significantly higher than those of castor oil cake-(15.9 mg g −1 ) [38], wheat straw-(32 mg g −1 ) [32], sugar cane-(87 mg g −1 ), orange peel-(28 mg g −1 ) [3] and plum stone-(179 mg g −1 ) [32] derived biochars, and close to those of grape stalk-(273 mg g −1 ) [32] and Alternanthera philoxeroides-(257 mg g −1 ) [40] derived biochars. However, CS-BC and MS-BC showed the lower Pb(II) removal potential as compared to grape husk-(595 mg g −1 ) [24] and waste-art-paper-(1200 mg g −1 and 1500 mg g −1 ) [39] derived biochars. According to the highest adsorption capacity of Pb(II) calculated by the Langmuir isotherm model (table 2), the BET surface area of the biochars did not seem to be the key factor controlling the adsorption of Pb(II). Chemisorption might be the controlling mechanism during Pb(II) adsorption on these biochars. At room temperature, the Pb(II) removal performance in aqueous solutions was nearly similar for both CS-BC and MS-BC. However, considering the yield of the biochar, the productivity of corn straw (44.4%) was notably higher than that of municipal sewage (26.3%). Thus, CS-BC has a big potential for the removal of Pb(II) from aqueous environment based on an economical point of view.

Conclusion
CS-BC and MS-BC both showed good potential for the removal of Pb(II) from aqueous solutions. The adsorbents loading, pH, temperature of the solutions and the functional groups on the biochar significantly influenced the Pb(II) adsorption performance on these biochars. CS-BC and MS-BC showed a maximum Pb(II) adsorption capacity of 352 and 387 mg g −1 , respectively, at pH = 7.0, 0.2 g l −1 biochar loading, 0.01 mol l −1 anion strength and 328 K. Unlike PO 3− 4 coexisting NO − 3 and CO 2− 3 in aqueous solutions did not significantly influence the adsorption of Pb(II). The adsorption experimental data were well fitted with Langmuir isotherm and pseudo-second-order kinetic models. Electrostatic interaction and surface complexation and coordination were suggested as Pb(II) adsorption mechanisms on biochars. The adsorption of Pb(II) over CS-BC and MS-BC was mainly carried out by chemisorption.