Abstract
A novel bentonite-biochar (APB) composite was prepared by incorporating bentonite (BE) with Alternanthera philoxeroides (AP) biochar for the adsorptive removal of Cd(II) from aqueous media. The APB and the pristine biochar (PB) prepared from the AP were produced at 300 °C under a nitrogen environment. The adsorption capabilities of the BE, PB, and APB were tested for the removal of Cd(II) from aqueous solution. The results showed that the pH substantially affected the adsorption of Cd(II) by the PB and APB. The adsorptive capacity of the Cd(II) onto the PB and APB gradually increased as the pH was increased to 6.0, and there was no significant change in adsorption as the pH was further increased to 8.0. The adsorption kinetic data of the PB and APB were fitted to a pseudo-second-order (PSO) adsorption kinetic model and an intraparticle diffusion (ID) model. The Freundlich model matched the experimental data better than the Langmuir model, indicating that the adsorption was heterogeneous. Thermodynamic study revealed that the adsorption was mainly physisorption, and the adsorption process was endothermic and spontaneous, while the orderliness of all adsorption systems decreased. The results demonstrated that the APB was an effective adsorbent for the removal of Cd(II) from aqueous media.
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Removal of Cd(II) from Aqueous Solution by Clay-biochar Composite Prepared from Alternanthera philoxeroides and Bentonite
Yande Jing,a,b Yongqiang Cao,a,b Qianqian Yang,a,b and Xuan Wang a,b
A novel bentonite-biochar (APB) composite was prepared by incorporating bentonite (BE) with Alternanthera philoxeroides (AP) biochar for the adsorptive removal of Cd(II) from aqueous media. The APB and the pristine biochar (PB) prepared from the AP were produced at 300 °C under a nitrogen environment. The adsorption capabilities of the BE, PB, and APB were tested for the removal of Cd(II) from aqueous solution. The results showed that the pH substantially affected the adsorption of Cd(II) by the PB and APB. The adsorptive capacity of the Cd(II) onto the PB and APB gradually increased as the pH was increased to 6.0, and there was no significant change in adsorption as the pH was further increased to 8.0. The adsorption kinetic data of the PB and APB were fitted to a pseudo-second-order (PSO) adsorption kinetic model and an intraparticle diffusion (ID) model. The Freundlich model matched the experimental data better than the Langmuir model, indicating that the adsorption was heterogeneous. Thermodynamic study revealed that the adsorption was mainly physisorption, and the adsorption process was endothermic and spontaneous, while the orderliness of all adsorption systems decreased. The results demonstrated that the APB was an effective adsorbent for the removal of Cd(II) from aqueous media.
Keywords: Clay-biochar; Cd(II); Aqueous solution; Removal
Contact information: a: College of Geography and Tourism, Qufu Normal University, 80 Yantai Road, Rizhao, Shandong, 276826, China; b: Key Laboratory of Nansi Lake Wetland Ecological and Environmental Protection in Universities of Shandong, 57 Jingxuan Road, Jining, Shandong, 273165, China; *Corresponding author: jingyande@163.com
INTRODUCTION
Cadmium(II) (Cd(II)) pollution has become a severe environmental problem around the world. More than 20,000 tons of Cd(II) enters the environment every year (Hayat et al. 2019). The concentration of Cd(II) in heavily polluted waters in China is as great as 4500 µg/L, while the national safety standard for Cd(II) is 0.1 mg/L (GB 8978-1996). Cadmium(II) and its compounds are soluble in water and have high fat solubility, bioaccumulation, and toxicity. They can enter the food chain and ultimately endanger human life and health.
Biochar has a well-developed porous structure. It has good adsorption properties for the removal of heavy metal ions and organic pollutants in water and soil (Ahmad et al. 2014; Cao et al. 2019). Using biochar to treat heavy-metal-polluted water has attracted wide interest over the past few decades. Alternanthera philoxeroides (AP), also known as revolutionary grass, water peanut, or lotus, is an invasive and harmful weed in China’s water and terrestrial ecosystem (Holm et al. 1977). It has strong resistance, wide adaptability, and rapid reproduction, and its invasion has caused profound harm to the environment (Li and Liu 2002). Currently, harvested AP is often treated by natural decomposition, which causes secondary environmental problems by releasing pathogens and methane (Xiao et al. 2009). However, conversion of the AP into biochar for remediation of heavy-metal-contaminated wastewater can utilize this harmful weed as a resource and avoid secondary pollution, while also expanding the material types of biochar and yielding economic benefits. To the authors’ knowledge, little research has been performed on using biochars from AP for the treatment of contaminated media. Only Huang et al. (2016) and Yang et al. (2014) have made preliminary explorations in this respect. Therefore, it will be of great ecological, environmental, and economic benefit to apply biochar prepared from AP to the remediation of heavy metals in wastewater.
At present, methods of removing Cd(II) from wastewater include extraction, membrane separation, ion exchange, and adsorption, among others (Gupta et al. 2009). The central pursuit of the adsorption method is seeking high-performance adsorbents. Recently, porous minerals (including zeolite, bentonite (BE), kaolin, and diatomite) have shown considerable advantages in the preparation of high-performance adsorbents (Wang et al. 2015a). Bentonite is a natural, layered, silicate clay mineral with a large specific surface area, excellent adsorption performance, and ion exchange capacity. It can be used for the remediation of heavy metal contamination with good results (Soetaredjo et al. 2017). However, BE tends to expand, disperse, and suspend in aqueous solution, and the effect of solid-liquid separation is poor, so it is limited in the treatment of wastewater (Luo et al. 2014). To improve the shortcomings of BE, some researchers have prepared a BE-based composite adsorbent and have used it to adsorb heavy metals in water (Fosso-Kankeu et al. 2017).
Clay-biochar composite adsorbents have been prepared in previous studies in attempts to increase the amount of pollutants that can be adsorbed compared to the pure adsorbents (Fosso-Kankeu et al. 2017). Bentonite-biochar composites have shown reasonably good results in previous studies (Laysandra et al. 2018). Previous research (Lee et al. 2013; Hao et al. 2017) has shown that the nature of the biomass remarkably affects the physicochemical composition of the biochar derived from it. Evidently, a replicate of the above composite with different biomass will exhibit different properties and affinities toward metals. To the authors’ knowledge, no studies have examined the application of composite adsorbents from AP and BE for the remediation of Cd(II)-contaminated wastewater. In this study, clay-biochar (APB) composites derived from the AP and BE were prepared at 300 °C, and the obtained composite adsorbents were evaluated for their ability to remove Cd(II) from aqueous solution. Equilibrium and kinetic experiments of Cd(II) adsorption with BE, pristine biochar (PB), and APB were performed. Furthermore, Fourier-transform infrared spectroscopy (FTIR), X-ray diffraction (XRD), and scanning electron microscopy (SEM) were used to characterize the structural differences in the BE, PB, and APB. The effects of pH and temperature were studied. Finally, the mechanisms of Cd(II) removal by the BE, PB, and APB were explored and discussed.
EXPERIMENTAL
Materials
The AP was collected from a pond in Rizhao, Shandong, China (35° 30’ N, 119° 16’ E). The collected material was cleaned with tap water and then cleaned with deionized water approximately three times. It was placed in a clean and ventilated location for natural air drying, washed, and oven-dried at 80 °C. Then, it was crushed by a grinder and sieved through a 100-mesh screen for subsequent experiments.
The calcium-based BE was purchased from Henan Yixiang Material Co., Ltd. (Henan, China). After simple physical purification, the calcium-based BE was screened using a 0.15-mm sieve. Reagents were purchased from Tianjin Hengxing Chemical Reagent Co., Ltd. (Tianjin, China). All other chemicals used in these experiments were analytical grade. All solutions were prepared with deionized (DI) water.
Clay-biochar composite preparation
The APB composite was prepared by adapting the procedure used by Yao et al. (2014). First, 2 g of the BE powder was dissolved in 500 mL of DI water and treated by sonication (KQ-100V; Kunshan Ultrasonic Instruments Co., Ltd., Kunshan, Jiangsu, China) of the mixture for 30 min to stabilize the BE suspension. Then, 10 g of the AP was weighed, immersed in the BE suspension, and stirred for 2 h. Finally, the feedstocks were separated from the mixture and oven-dried at 80 °C.
The feedstocks were air-dried and ground, packed into a closed ceramic crucibles, and put into a preheated (75 C) muffle furnace (KSY-12D-16, Longkou City Factory Furnace, Wuhan, China). The feedstocks were decomposed for 2 h at 300 C in oxygen-limited conditions with the inert atmosphere achieved with an N2 flow rate of 0.5 L/min. The heating rate was 15 °C/min. After being cooled to room temperature, the solids were cleaned several times with DI water to remove surface impurities and were oven-dried at 80 °C (Wang et al. 2015b). The PB, without BE modification, underwent the same pyrolysis process. The obtained adsorbents were ground through a 0.15-mm sieve and stored for later use. The BE-modified biochar and the unmodified biochar were denoted “APB” and “PB,” respectively.
Methods
Adsorbent characterization
The surface morphologies and pore structures of the adsorbents were determined using SEM (JSM-5600LV; JEOL Ltd., Tokyo, Japan) at 1000× magnification coupled with an energy dispersive spectroscopy (EDS) instrument (Oxford model 7582, ISIS, USA). Fourier-transform infrared spectroscopy (Tensor 27; Bruker, Karlsruhe, Germany) was used to analyze the surface functional groups of the adsorbents. The adsorbents were ground and mixed with KBr wafers (0.5%). At a 4 cm-1 resolution, the spectra were obtained over the range of 500 cm-1 to 4000 cm-1 with 64 scans. A CHN elemental analyzer (Vario EL Cube; Elementar Analysensysteme GmbH, Langenselbold, Germany) was used to determine the elemental composition (C, H, N, S) of the adsorbents, and the O content was obtained by a mass balance of ash content and elemental contents.
The pH values of the adsorbents were determined as follows: First, 2.5 g of the adsorbent was weighed and added into 50 mL of deionized water. The adsorbent was heated in airtight conditions, boiled for 5 min, and filtered, discarding 5 mL of primary filtrate. The pH of the residual liquid after cooling was measured using a pH meter.
The ash contents of the adsorbents were determined as follows: First, 1.0 g of the adsorbent was weighed and burned in a muffle furnace at 650 C for 1 hour in aerobic conditions. The ash content was calculated according to the quality of the adsorbents before and after burning.
Adsorption experiments
The adsorption behavior was modeled by three methods: the isotherm model, the kinetic model, and the thermodynamic model. Before applying these models, the adsorption capacity was first determined using Eq. 1,
(1)
where qt is the amount of Cd(II) adsorbed per unit mass of adsorbent (mg/g), C0 is the initial concentration of Cd(II) in the solution (mg/L), Ct is the Cd(II) concentration (mg/L) at time t (with the equilibrium concentration also denoted Ce), m is the mass of the adsorbent (g), and V is the solution volume (L).
In the kinetic experiment, the adsorbent dosage was maintained at 0.5 g/L with a Cd(II) concentration of 30 mg/L. The pH of the system was maintained in the range of 7.0 to 8.0. A shaking speed of 200 rpm was maintained during the kinetic experiment. The residence time was varied (5 min, 10 min, 20 min, 30 min, 40 min, 1 h, 1.5 h, 3 h, 6 h, 9 h, 15 h, and 24 h). After each run, the pH values of the suspensions were measured. Then, the concentrations of Cd(II) in the suspensions were determined using an atomic absorption spectrophotometer (GFA-7000A; Shimadzu, Kyoto, Japan) after the suspensions were filtered through a 0.45-μm-filter membrane. Three replications were performed for each contact time. Kinetic modeling was performed using the Origin 9.1 software package (OriginLab, Northampton, MA, USA).
To investigate the adsorption processes of the BE, PB, and APB for Cd(II), different sorption kinetic models were applied to fit the experimental data. These were the pseudo-first-order (PFO) kinetic model (Eq. 2), the pseudo-second-order (PSO) kinetic model (Eq. 3), and the intraparticle diffusion (ID) model (Eq. 4) (Bazrafshan et al. 2015),
where qe and qt are the amounts of Cd(II) adsorbed onto the adsorbent at equilibrium and at time t (mg/g), respectively, k1 (min−1) and k2 (g/(mg·min)) are the rate constants of the PFO and PSO adsorption models, respectively, kp is the ID rate constant (mg/(g·min1/2)), and c is the linear intercept.
Adsorption isotherms were determined for different Cd(II) concentrations (20 mg/L, 30 mg/L, 40 mg/L, 50 mg/L, 60 mg/L, 80 mg/L, and 100 mg/L) with a defined pH (pH 7 to 8), adsorbent dosage (0.5 g/L), and residence time (24 h). Other experimental conditions remained similar to the edge and kinetic experiments. The pH values of the suspensions were measured, and the Cd(II) concentrations of the suspensions were analyzed, as described above. Isotherm data for the adsorption of Cd(II) onto the BE, PB, and APB were modeled using the Langmuir (Eq. 5) and Freundlich (Eq. 6) isotherm models,
where Kf (mg/g)·(mg/L)−n and KL (L/mg) are the Freundlich and Langmuir affinity parameters, respectively, qe is the equilibrium adsorption capacity (mg/g), Ce is the equilibrium liquid phase concentration (mg/L), qm is the maximum adsorption capacity in the Langmuir model (mg/g), and n is a Freundlich constant related to adsorption.
To obtain a complete description of the adsorption of Cd(II) onto the BE, PB, and APB, several thermodynamic properties of the adsorption system were assessed. The thermodynamic feasibility of the adsorption of Cd(II) onto the BE, PB, and APB was assessed through the standard Gibbs free energy change (ΔG°, kJ/mol), standard enthalpy change (ΔH°, kJ/mol), and standard entropy change (ΔS°, kJ/(mol·K)). The standard Gibbs free energy change was determined by Eq. 7,
where Ko is a thermodynamic distribution coefficient. The value of Ko was obtained from Eq. 8 (Khan and Singh 1987), as shown in Eq. 9. The relation between the thermodynamic distribution coefficient and ΔH° and ΔS° is given by Eq. 9,
(9)
where R is the gas constant (8.314 J/(mol·K)), T is the absolute temperature (K), and Ko is the equilibrium constant (m/M). The values of ΔH° and ΔS° could be calculated from the slope and intercept of the linear plot of ln Ko (Ko = qe/Ce) vs. 1/T.
pH-dependent Experiment
The effects of pH changes on Cd(II) adsorption by the BE, PB, and APB were studied by adjusting the pH of 40 mL of Cd(II) solution (30 mg/L) with 0.1 M HNO3 or NaOH in the pH range of 2.0 to 8.0. Then, 0.02 g of adsorbent was added into a 100-mL centrifugal tube. The concentration of Cd(II) was determined using the atomic absorption spectrophotometer after 24 h by shaking the mixture at 200 rpm at 25 °C. The adsorption capacity of Cd(II) onto the adsorbents at different pH levels was calculated as described above.
RESULTS AND DISCUSSION
Characterization of the Adsorbents
The basic properties of the adsorbents are listed in Table 1. The ash contents of the PB and APB were 14.3% and 19.9%, respectively. The yields of the PB and APB were 50.4% and 56.0%, respectively. The ash content and yield of the APB were greater than those of the PB. This result might have been because the APB had the BE particles attached to the surface of the PB, thus increasing the content of mineral components, such as Si, Mg, and Al in the APB, which increased the content of the APB after pyrolysis (Xu et al. 2013). As shown by the elemental compositions of the PB and APB, the most abundant element in the PB and APB was C, followed by O. Such results are consistent with other work (Cao et al. 2019). The H/C, O/C, and (O + N) / C ratios are often used to characterize the aromaticity and carbonization degree of adsorbents and the number and polarity of surface oxygen-containing functional groups (Ren et al. 2018). As shown in Table 1, the aromaticity of and the number and polarity of oxygen-containing functional groups in the PB were greater than or equal to those of the APB at the same pyrolysis temperature.
The pH values of the PB and APB were all greater than 7, indicating weak alkalinity and mainly being because the PB and APB used AP as a raw material. Organic acids and phenolic substances were formed due to the decomposition of hemicellulose and cellulose in the AP, which resulted in the pH of adsorbents being near neutrality. Furthermore, the pH of the APB was slightly lower than that of PB. This result was consistent with prior research (Premarathna et al. 2019) and was due to the inhibition of the BE adhering to the surface of the AP.
Table 1. Yield, pH, Ash Content, Elemental Composition, and Atomic Ratios
The SEM images and EDS (energy-dispersive X-ray spectroscopy) analysis of the adsorbents are shown in Fig. 1.
Fig. 1. SEM images and EDS spectra of (a) APB and (b) PB
The SEM results showed that the surface morphologies of the adsorbents varied remarkably between the different types. The PB had a completely tubular or sheet structure with a distinct pore or groove structure on its surface (Wang et al. 2019). The surface morphology of the APB was roughly tubular or of an irregular lamellar structure. The coverage of the BE particles on the surface of the AP biochar was evidenced by the EDS analysis. The BE particles could provide more adsorption sites for the APB. As shown in Fig. 1, compared with the PB, the EDS spectrum of the APB surface showed that the peak values of Si, Ca, Mg, and Al were relatively high, all of which are typical of the elemental composition of montmorillonite, the main component of the BE. This result also shows that the BE was successfully loaded onto the AP biochar.
Fourier-transform infrared spectroscopy was used to qualitatively determinate the functional groups of the BE, PB, and APB. Figure 2 shows the FTIR spectra of the adsorbents. As shown, the types and richness of oxygen-containing functional groups differed greatly among the adsorbents. The stretching vibration absorption peaks of Si-O-Si and Si-O existing in the APB composites were similar to those of the BE, which further indicated that the BE had been loaded onto the surface of the AP biochar.
Fig. 2. FTIR spectra of the PB, APB, and BE
The BE exhibited stretching vibration absorption peaks of hydroxyl (-OH) and hydrogen bonds (-OH) at 3625 cm-1 and 3417 cm-1, respectively (Paluszkiewicz et al. 2008). The absorption peak near 2881 cm-1 was caused by the absorption vibration of -CH. The stretching vibration at 1763 cm-1 was C=O, and the absorption peaks near 1628 cm-1 and 1426 cm-1 were -NH bending vibrations. The absorption peaks at 1042 cm-1 and 531 cm-1 were caused by Si-O stretching vibration and Si-O-Al deformation vibration, respectively, while the absorption peak at 775 cm-1 was caused by Si-O-Si stretching vibration (Alabarse et al. 2011).
The broad absorption peaks of the PB near 3430 cm-1 and 3244 cm-1 were caused by the stretching vibrations of phenolic hydroxyls (-OH) (Yu et al. 2018; Khan et al. 2010). Many absorption peaks near 2839 cm-1 to 2987 cm-1 were produced by the stretching vibration of aliphatic -CH2 (Zhang et al. 2011). The C-O absorption peak of carboxyl groups was at 1394 cm-1, the C=C stretching vibration peak of aromatic rings was at 1610 cm-1 (Zhao et al. 2013), and the C-O-C stretching vibration peak of ether in cellulose, hemicellulose, or lignin was at 1047 cm-1 (Premarathna et al. 2019). The bending vibration absorption peaks of aromatic C-H were mainly located in the wavenumber band from 500 cm-1 to 779 cm-1. The absorption peaks near 1610 cm-1, 1394 cm-1, and 1047 cm-1 decreased, which indicated that the number of carboxyl, ketone, and ether functional groups was relatively small.
The APB had a stretching vibration absorption peak of -CH near 2883 cm-1, which was similar to that of the PB near this band, indicating that all adsorbents had aliphatic functional groups. The peaks near 1763 cm-1 and 1610 cm-1 corresponded to the peaks of the PB, which were the stretching vibration of C=O and the stretching vibration of C=C on aromatic rings. Near 1315 cm-1 was the stretching vibration peak of C-O, and near 1029 cm-1 was the Si-O-Si stretching vibration absorption peak (Ünlü et al. 2012). At 780 cm-1, the Si-O-Si stretching vibration absorption peak was similar to that of the BE. These characteristics indicated that the BE had successfully adhered to the surface of the AP biochar.
pH Edge Experiments
The pH of a solution is a major factor that controls adsorption onto adsorbents (Li et al. 2017; Premarathna et al. 2019). The solution pH affects not only the surface charge of the adsorbents but also the forms and ionization degrees of heavy metals in solution, along with the adsorption rate and degree (Chang et al. 2016). When the pH is less than 8.0, Cd(II) mainly exists in the form of free Cd(II) and a small amount of Cd(OH)+ ions in the solution. Greater pH of the solution (pH ≥ 8) will cause the precipitation of Cd(II) and interfere with the actual adsorption of Cd(II) onto the adsorbents (Chen et al. 2011). Therefore, this study examined solution pH in the range of 2 to 8 to test the adsorption of Cd(II) onto the adsorbents. As shown in Fig. 3, the initial pH of the solution had a substantial influence on the adsorption of the Cd(II).
Fig. 3. Effects of initial solution pH on Cd(II) adsorption (Data are the mean ± standard deviation based on duplicate measurements).
Overall, the adsorbed amount of Cd(II) increased rapidly at first and then plateaued with increasing pH. The greatest Cd(II) adsorption by the PB (37.61 mg/g) and APB (44.3 mg/g) was observed at a pH of 6.0. Compared to the PB and APB, the BE showed a similar trend in adsorption capacities in the pH range of 2.0 to 8.0, and the maximum adsorption capacity was similarly obtained at a pH of 6.0. There are three possible reasons for the decreased Cd(II) adsorption capacities at low pH values. First, a large number of H+ ions and Cd(II) in the solution competed for limited binding sites, which reduced the adsorption of Cd(II). Second, the increased solubility of crystal minerals in the adsorbents resulted in the release of many cations, such as K+, Ca2+, Mg2+, which competed with Cd(II) in the solution for limited adsorption points. Third, anions, such as CO32- and PO43-, were difficult to precipitate with Cd(II) in solution. As the solution pH increased, the amount of H+ in the solution decreased, and the cations (such as K+, Ca2+, and Mg2+) released by the adsorbent material also decreased, which reduced the competitive adsorption with Cd(II). Moreover, Cd(II) could easily precipitate with anions such as CO32- and PO43-; thus, the adsorption capacity of Cd(II) onto the adsorbents increased. Moreover, the negative charge density on the surfaces of the adsorbents increased with the increased pH, which was conducive to weakening the electrostatic repulsion of Cd(II) and thus increasing the adsorption capacity (Rivera-Utrilla et al. 2013).
Kinetic Experiments
Figure 4 shows the adsorption kinetics curve and models of the Cd(II) adsorption.
Fig. 4. Kinetic curve and fitted curves of Cd(II) adsorption: (a) kinetics curve, (b) PFO model, (c) PSO model, and (d) ID model
The adsorption process could be divided into two distinct stages. The adsorption capacity of Cd(II) onto the adsorbents increased rapidly with increasing time in the early stage and then plateaued. At 360 min, the adsorption capacities of the BE, PB, and APB to Cd(II) in solution reached 28.2 mg/g, 26.6 mg/g, and 35.3 mg/g, respectively, accounting for more than 95% of the adsorption capacity at 1440 min, basically reaching the equilibrium state of adsorption. The decrease of the active adsorption sites on the adsorbent surfaces caused the decrease of the adsorption rate as time progressed, which caused the adsorption of Cd(II) to diffuse into the adsorbent material.
To better explore the adsorption process and mechanism of Cd(II) and the different adsorbents, based on the analysis of the adsorption kinetics curve, the experimental data were fitted by a PFO kinetic model (Fig. 4b) and a PSO kinetic model (Fig. 4c). The parameters of the PFO and PSO models obtained from the fitting of the experimental data are summarized in Table 2. Both of these models could represent the adsorption kinetic data well, according to the squared correlation coefficients (R2 > 0.90). The R2 values of the PSO kinetic model were all greater than those of the PFO kinetic model. The theoretical equilibrium adsorption capacities (qe, cal) of the PSO kinetic model were 27.3 mg/g (PB), 36.9 mg/g (APB), and 28.9 mg/g (BE), which were close to the actual equilibrium adsorption quantities (qe, exp) of 27.4 mg/g (PB), 36.9 mg/g (APB), and 28.8 mg/g (BE). The PSO model had smaller deviations than the PFO model, which indicated that the PSO kinetic model could better describe the Cd(II) sorption kinetics. It has been shown recently that good fits to the PSO model are generally consistent with diffusion as the rate-limiting process (Hubbe et al. 2019).
The parameters k1 and k2 reflect the time required for the adsorption process to reach equilibrium in the PFO and PSO kinetic models, respectively (Plazinski et al. 2009). These parameters strongly depended on the applied operating conditions, such as temperature, initial concentration, solution pH, and agitation rate. The adsorption rate constant k2 was greater for PB (0.0072) than for APB (0.0026), and the time required for the PB to reach equilibrium was less than that for the APB. A possible reason was that it took more time and process for Cd(II) to adsorb on the APB due to the BE loading on the surface of the APB.
Table 2. Kinetic Parameters Obtained from PFO and PSO Kinetic Models of Cd(II) Adsorption onto the PB, APB, and BE
To determine the actual rate-controlling steps of Cd(II) onto the PB and APB, the ID model was used to fit the experimental data, and the curve in Fig. 4d and the parameters in Table 3 were obtained. The whole adsorption process could be divided into two stages: The first stage (0 min to 60 min) was the process of Cd(II) diffusion onto the surface of the PB and APB, and kp1 was greater than kp2, indicating that the boundary diffusion process was fast, and the external surface adsorption was dominant. In the second stage (70 min to 1440 min), the fitting curve plateaued, and kp2 was less than kp1, indicating that the adsorption of Cd(II) onto the PB and APB was a rate-controlling step in this stage. As shown in Table 3, the diffusion rate constant kp1 was greater than kp2. This result was mainly due to the decrease of Cd(II) concentration in the solution, which led to the decrease of the second stage diffusion rate and the final equilibrium state of Cd(II) adsorption. Meanwhile, the adsorption capacity on the external surface of the adsorbents tended to be saturated in the second stage. With the increase toward the adsorption capacity of the inner surface, the diffusion resistance of Cd(II) increased gradually, which led to a decrease in the diffusion rate. The C value in Table 3 indicated the thickness of the adsorbent boundary layer. The C value decreased with the increase of heterogeneity and hydrophilic groups on the surfaces of the adsorbents. As shown, C1 was less than C2, indicating that the adsorption effect of Cd(II) onto the adsorbents’ boundary layers was much greater in the second stage than in the first stage. Figure 4d shows that the adsorption process of the PB and APB for Cd(II) corresponded to multistage linear fitting. None of the straight lines passed through the origin but with positive intercept, indicating that the adsorption process was complicated and had a specific internal diffusion, which might be one of speed-controlling steps. There might be other unknown speed-controlling steps (Ma et al. 2011). The actual sorption process included surface adsorption and liquid membrane diffusion, which jointly controlled the adsorption reaction rate (Kara and Tuncel 2011).
Table 3. Parameters of the ID Model for Cd(II) Adsorption
Isotherm Experiments
The isothermal adsorption experiments of Cd(II) onto the adsorbents in solution were performed at 288 K, 298 K, and 308 K. Figure 5 depicts the experimental data fitted by the Freundlich and Langmuir isothermal adsorption models. In Fig. 5, the adsorption experimental data are indicated with symbols, while the solid lines represent the theoretical data from the models. The temperature-dependent parameters obtained from the fitting of the experimental data are summarized in Table 4.
Figure 5 shows that with increasing equilibrium concentration of Cd(II) (Ce), the equilibrium adsorption capacity (qe) gradually increased, and the growth rate decreased. Furthermore, the actual adsorption capacity of the BE, PB, and APB increased with increasing temperature, which indicated that high temperature was beneficial to the adsorption of Cd(II) onto the BE, PB, and APB, and the adsorption was spontaneous.
Table 4 shows that Freundlich isothermal adsorption model (R2 = 0.9360 to 0.9932) and the Langmuir isothermal adsorption model (R2 = 0.8202 to 0.9650) fitted the experimental data of Cd(II) adsorption onto the BE, PB, and APB well. Specifically, the fitting coefficients of the Freundlich model (R2 = 0.9360 to 0.9932) were greater than those of Langmuir model (R2 = 0.8202 to 0.9650), which indicated that the Freundlich model could better describe the adsorptive behavior of the Cd(II) onto the adsorbents.
Fig. 5. Adsorption isotherms of Cd(II) at different temperatures
Table 4. Parameters of the Isothermal Sorption Models for Cd(II) Adsorption
The adsorption capacity would continue to increase with the increase of Cd(II) concentration, while there were some inhomogeneities in adsorption. This was consistent with the results of Moyo et al. (2016). The non-linear index n in the Freundlich model was less than 1, which indicated that the distribution of Cd(II) adsorption sites on the adsorbents were heterogeneous (Fosso-Kankeu et al. 2017).
The parameter Kf measures the strength of the attraction and attachment of Cd(II) onto the surface of the adsorbents. The high Kf value indicates that the attraction force was strong. Table 4 shows that Kf increased with increasing temperature, indicating that the adsorptive force of the adsorbents increased with increasing temperature. At 298 K, the adsorption forces of the adsorbents were ordered as follows: APB (21.35) > BE (20.57) > PB (16.18). According to the parameter qm of the Langmuir model in Table 4, the theoretical maximum adsorption capacities (qm) for the adsorption of Cd(II) onto the adsorbents were ordered as follows: 111.9 mg/g (APB) > 108.0 mg/g (BE) > 100.4 mg/g (PB), which was the same order as that of the Kf values. The APB had better adsorptive force or capacity for Cd(II) mainly because the surface of the APB contained various functional groups from the bentonite or the AP biochar, and these functional groups enhanced the attraction force of the adsorbents. Furthermore, the loading of the BE onto the APB composite adsorbent increased the adsorptive sites and enhanced ion exchange.
Thermodynamic Experiments
The environmental temperature was one of the main factors affecting adsorption. The experimental data of the Cd(II) adsorption at 288 K, 298 K, and 308 K were analyzed by thermodynamic correlation (Fig. 6). The thermodynamic parameters KD, ΔG°, ΔH°, and ΔS° are summarized in Table 5.
Fig. 6. Effect of temperature on coefficients
The spontaneity of the adsorption process was measured by the standard Gibbs free energy (ΔG°). The negative value of ΔG° indicates that the adsorption process was spontaneous; the large negative value of ΔG° indicates that the system had high adsorption affinity. Table 5 shows that the values of ΔG° for the different adsorbents were negative at all three temperatures, which indicated that the adsorbents adsorbed Cd(II) spontaneously. The values of ΔG° decreased with increasing temperature, which indicated that increased temperature was beneficial to the spontaneous adsorption. Generally, values of ΔG° in the range of -20 kJ/mol to 0 kJ/mol indicate that physisorption is the primary adsorption process, while chemisorption is the main adsorption process in the range of -400 kJ/mol to -80 kJ/mol (Yu et al. 2004). In this study, ΔG° was in the range of -5.50 kJ/mol to -4.07 kJ/mol, indicating that the adsorption process of Cd(II) onto the adsorbents was mainly physisorption.
A positive value of ΔH° indicates that an adsorption process is endothermic, and the uptake of adsorbate molecules increases with increasing temperature. A negative value of ΔH° indicates that the process is exothermic, and the increased temperature has an adverse effect on the number of adsorbed molecules. The enthalpy change (ΔH°) was positive (Table 5), indicating that the adsorption of Cd(II) by the adsorbents at 288 K to 308 K was an endothermic reaction, and the uptake of Cd(II) onto the adsorbents increased with increasing temperature.
The standard entropy change (ΔS°) measures the randomness of the adsorption system. A positive value of ΔS° shows increased randomness of the interface between adsorbents and adsorbate molecules (Kara and Tuncel 2011; Fil et al. 2012), while a negative value of ΔS° indicates an increase of orderliness of the adsorption system. With greater order of the system, fewer adsorbate molecules are adsorbed onto the surface of the adsorbents. The values of ΔS° were positive for all systems (Table 5), indicating a decrease in the orderliness of all the adsorption systems. The lowest value of ΔS° was for BE. This might be because the BE was a clay mineral, which had an ordered crystalline layered structure composed of polymeric sheets of tetrahedral SiO4 linked into sheets of octahedral (Al, Mg, Fe)(O, OH)6 (Ismadji et al. 2015).
Table 5. Thermodynamic Parameters of Cd(II) Sorption
CONCLUSIONS
- A novel clay-biochar composite was prepared by thermal treatment of bentonite (BE) and Alternanthera philoxeroides (AP) at 300 °C under a nitrogen environment. The SEM and EDS analyses showed that the bentonite-biochar (APB) produced more fragmented substances and had a more developed pore and groove structures than the pristine biochar (PB). The FTIR analysis showed that the stretching vibration absorption peaks of Si-O-Si and Si-O in the APB composites were similar to those in the BE, which indicated that the BE had been loaded onto the surface of the AP biochar.
- The pH markedly affected the adsorption of the Cd(II) by the PB and APB. The adsorptive capacity of the Cd(II) onto the PB and APB gradually increased with increasing pH in the range of 2.0 to 8.0 and reached equilibrium at a pH of 6.0. The adsorption kinetic data of the PB and APB fitted the PSO adsorption kinetic model and the ID model, which indicated that the time required for the PB to reach equilibrium was less than that for the APB and the second stage of the adsorption process of Cd(II) onto the PB and APB was the rate-controlling step. The rate of the adsorption reaction might have been controlled by surface adsorption, liquid membrane diffusion, and internal diffusion. The Freundlich model fitted the adsorption isotherm data for all the adsorbents better than the Langmuir model, indicating that the adsorption was a multilayer adsorption process involving, and the distribution of the adsorptive sites of Cd(II) onto the adsorbents was heterogeneous. The thermodynamic study indicated that the adsorption of Cd(II) onto the adsorbents was mainly physisorption, and the adsorption was an endothermic and spontaneous process, while the orderliness of the adsorption system decreased.
- The APB showed much higher Cd(II) adsorption capacity than the PB and BE. In summary, both waste reuse and Cd(II) removal could been achieved by the APB. The APB could be used as an effective and eco-friendly adsorbent to enhance the removal of Cd(II) from aqueous solution, and it can also provide a theoretical basis for applications in heavy metal pollution remediation from other media.
ACKNOWLEDGMENTS
This research was supported by the Natural Science Foundation of Shandong Province (No. ZR2013DM005) and the Humanities and Social Science General Project of the Ministry of Education (No. 15YJAZH027). The authors extend sincere gratitude to Yongqiang Cao for help in setting up and running the experiments.
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Article submitted: August 3, 2019; Peer review completed: November 3, 2019; Revised version received: November 30, 2019; Accepted: December 1, 2019; Published: December 4, 2019.
DOI: 10.15376/biores.15.1.598-615