chitosan modified biochar microspheres derived from Thalia dealbata

Bioresource Technology 218 (2016) 1123–1132

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Removal of phosphate from aqueous solution using magnesiumalginate/chitosan modified biochar microspheres derived from Thalia dealbata Xiaoqiang Cui a, Xi Dai a, Kiran Yasmin Khan a, Tingqiang Li a, Xiaoe Yang a,⇑, Zhenli He b a Ministry of Education Key Laboratory of Environmental Remediation and Ecological Health, College of Environmental and Resource Sciences, Zhejiang University, Hangzhou 310058, China b Indian River Research and Education Center, Institute of Food and Agricultural Sciences, University of Florida, Fort Pierce, FL 34945, USA

h i g h l i g h t s

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


Five types of modified biochars were

synthesized for phosphate remediation.
The introduction of MgCl2 substantially increased surface area of biochar.
MgCl2-alginate biochar microspheres had the highest phosphate sorption capacity.
Phosphate sorption on biochars fits a pseudo second order and Langmuir model.
Precipitation and ligand exchange were the possible sorption mechanisms.

a r t i c l e

i n f o

Article history: Received 1 June 2016 Received in revised form 17 July 2016 Accepted 19 July 2016 Available online 21 July 2016 Keywords: Biochar Phosphate Sorption characteristics Microspheres Modification

a b s t r a c t The objective of this study was to determine the feasibility of using magnesium-alginate/chitosan modified biochar microspheres to enhance removal of phosphate from aqueous solution. The introduction of MgCl2 substantially increased surface area of biochar (116.2 m2 g1), and both granulation with alginate/chitosan and modification with magnesium improved phosphate sorption on the biochars. Phosphate sorption on the biochars could be well described by a simple Langmuir model, and the MgCl2-alginate modified biochar microspheres exhibited the highest phosphate sorption capacity (up to 46.56 mg g1). The pseudo second order kinetic model better fitted the kinetic data, and both the Yoon-Nelson and Thomas models were superior to other models in describing phosphate dynamic sorption. Precipitation with minerals and ligand exchange were the possible mechanisms of phosphate sorption on the modified biochars. These results imply that MgCl2-alginate modified biochar microspheres have potential as a green cost-effective sorbent for remediating P contaminated water environment. Ó 2016 Elsevier Ltd. All rights reserved.

1. Introduction

⇑ Corresponding author. E-mail address: [email protected] (X. Yang). http://dx.doi.org/10.1016/j.biortech.2016.07.072 0960-8524/Ó 2016 Elsevier Ltd. All rights reserved.

Phosphorus (P) is regarded as an essential element in agriculture and industry, but excessive release of P into natural water may cause eutrophication, which not only imposes risk to the aquatic ecosystems but also results in economic loss (Acelas

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et al., 2015; Krishnan and Haridas, 2008; Yao et al., 2013). According to the national environment survey of China in 2014, eutrophic lakes and reservoirs accounted for 24.59% of the national total surface water area (MEPC, 2015). It is imperative to develop effective techniques to remove phosphate from aqueous solution. On the other hand, as a finite resource on earth, recycle of the removed phosphorus should be considered. To date, many methods have been developed including biological treatment, chemical precipitation and adsorption (Yao et al., 2013; Zhao et al., 2012). Among them, adsorption is considered to be more useful and economical, especially for the low P concentration environment (Ma et al., 2011). A variety of sorbents have been compared for their removal efficiency of P in aqueous solution, such as fly ash, crop residues, steel slags and biochars (Acelas et al., 2015; Krishnan and Haridas, 2008; Lu et al., 2009; Xiong et al., 2008; Yao et al., 2013). The low cost sorbents such as crop residues usually have a lower sorption capacity, while the cost of synthesized sorbents is often too high for field application. Hence, cost-effective sorbents are needed for enhancing P removal from eutrophic waters. Biochar is a solid material obtained from the thermochemical conversion of plant biomass in an oxygen-limited environment, which has received much attention for soil improvement, waste management, climate change mitigation and energy production. Recently, biochar has been widely used as a green sorbent for removal of contaminants such as organic pollutants, heavy metals and others (Ahmad et al., 2014). However, most of biochars have a lower sorption capacity for phosphate and other anionic pollutants than heavy metals. Sarkhot et al. (2013) reported that up to 0.24 mg g1 phosphate was removed from dairy manure by biochar prepared from mixed hardwood shavings. Hale et al. (2013) found that cacao shell and corn cob derived biochars had minimal sorption capacity for phosphate and nitrate. Jung et al. (2015) reported that the maximum phosphate sorption capacity of peanut shell derived biochar was only 6.79 mg g1. Apparently, the P removal capacity of biochars is related to the properties of raw materials. It may be necessary to modify plant biomass based biochars in order to effectively remove P in water. Moreover, most of currently used biochars are mostly derived from crop residues, with much less from wetland plants. Thalia dealbata, a typical emergent aquatic plant, is widely grown in the floating islands and constructed wetlands for purifying eutrophic water (Zhao et al., 2012), and the planting area are rapidly increasing in China. However, it is essential to periodically harvest plant biomass to avoid the return of nutrients back to water from decomposing plant litters. Consequently, disposal of a large quantity of plant biomass has become a challenge, and converting the biomass into biochar used as a sorbent is considered as an environmental and economic strategy (Cui et al., 2016). In order to improve the sorption capacity of biochars for phosphate, several surface modifications were compared in this study. In most of the previous investigations, the sorption capacity of biochars was usually evaluated with batch sorption experiments, as the powdered biochars are not suitable for dynamic sorption application. Limited studies have been performed on the dynamic sorption of phosphate on biochar particles, while the practical sorption process is usually operated using fixed beds at a large scale wastewater treatment. In addition, lack of collection and regeneration efficiency of powder biochar has limited its application. Therefore, development of biochar particles is urgently needed for dynamic sorption of phosphate via regeneration and repeated use (Do and Lee, 2013; Roh et al., 2015). Alginate and chitosan are nontoxic and inexpensive natural materials, which have been widely used as gelling agents to fabricate microcapsules or microspheres. Recently, alginate and chitosan have been used as natural polymer sorbents for the removal of heavy metal ions, including Pb2+, Cu2+, Zn2+, and Cd2+ (Do and Lee, 2013; Zhou

et al., 2013). However, to date, no study has reported the use of chitosan and alginate to obtain Mg-modified biochar microspheres derived from T. dealbata for enhancing their phosphate sorption capacity. Furthermore, the granular biochar microspheres are easily to collect and applied as a potential fertilizer after phosphate sorption process, which could be an ideal way to remove contaminant and recycle resource (Jung et al., 2016; Yao et al., 2013). In this study, Mg-alginate/chitosan modified biochars were synthesized and evaluated for their phosphate sorption capacity. The original biochar was obtained from T. dealbata through slow pyrolysis at 500 °C, and the modified biochars were prepared with Mg (II) salt solution, sodium alginate, and chitosan. After characterization of biochars, pH dependent soprtion, sorption kinetics and isotherms of phosphate on selected biochars were determined to compare their sorption capacity and understand sorption mechanisms of the modified biochars. Meanwhile, column experiments were conducted to examine the dynamics of phosphate sorption on the biochars in aqueous solution. 2. Materials and methods 2.1. Preparation of original and modified biochars T. dealbata sample was collected from the Ningbo constructed wetland, Zhejiang province, China. The washed and air-dried sample was dried at 85 °C for 24 h and sieved to <2.0 mm particles using a stainless grinding machine (Retsch MM400, Germany). The powdered biomass was pyrolyzed to 500 °C at a rate of 5 °C/ min and held at the peak temperature for 2 h before cooled to room temperature in a muffle furnace under N2 atmosphere, and the biochar (i.e. the solid residues from the pyrolysis) was referred to as TB. The MgCl2 modified biochar sample was obtained by pyrolyzing the powdered biomass immersed with 1 M MgCl2 solution under the same pyrolysis conditions, and identified as TBM. All of the powdery biochar samples were ground to pass through a 0.25-mm sieve prior to use. For the preparation of composite alginate-biochar microsphere samples, the biochar (10%) was incorporated with sodium alginate (2%), and this mixed solution was stirred for 30 min. Then the biochar-alginate suspension was dropped with a flow rate using a syringe into calcium chloride solution (3%) to get homogeneous microspheres. These microspheres were allowed to harden for 1 h and rinsed with deionized water at room temperature, and kept in the calcium chloride solution for 24 h at 4 °C. The alginate modified biochars were then washed with deionized water and oven-dried for 5 h at 60 °C. The composite alginate-biochar microsphere samples are hereafter referred to as TB-A and TBM-A, respectively. For the preparation of composite biochar-chitosan microsphere samples, 2 g of chitosan was dissolved in 100 mL acetic acid (2%). Then, 2 g of biochar was added and the mixture was stirred for 30 min and the biochar-chitosan suspension was dropped with a flow rate using a syringe into a 100 mL NaOH (3%) and kept in the solution for 12 h. The chitosan modified biochars were then washed with deionized water and oven-dried for 5 h at 60 °C. The composite biochar-chitosan microsphere samples are hereafter referred to as TB-C and TBM-C, respectively. A schematic illustration of the preparation process is depicted in Fig. S1. 2.2. Characterization The surface area of the biochars was characterized by N2 adsorption isotherms at 77 K with the Brunauer-Emmett-Teller (BET) method using a Quadrasorb Si-MP surface area analyzer. The morphology and microstructure analysis were conducted using a scanning electron microscope (SEM, FEI QUANTA FEG

X. Cui et al. / Bioresource Technology 218 (2016) 1123–1132

650), and surface element analysis was performed simultaneously with the SEM at the same surface locations using energy dispersive X-ray spectroscopy (EDS, EDAX Inc. Genesis XM). Elemental (C, H, N) analyses were conducted using a CHN Elemental Analyzer (Flash-EA112, Thermo Finnigan). A Fourier transform infrared spectrometer (FTIR, Nicolet 6700) was employed to collect spectra in the 400 and 4000 cm1 region with 50 scans being taken at 2 cm1 resolution to identify the surface functional groups. The mineral crystallographic structure and surface chemical composition were measured by X-ray diffraction (XRD) using a D8 Advance Bruker spectrometer. Zeta-potential was measured at different pH values with a potential analyzer (Malvern Nano-ZS90, England). Ash content was measured by heating the biochar samples at 750 °C for 5 h. The pH of biochars was measured by a pH meter with a mass/volume ratio of 1:20 for biochar to deionized water. 2.3. Batch sorption experiments Phosphate solutions were prepared by dissolving potassium dihydrogen phosphate (KH2PO4, certified A.C.S, Fisher Scientific) in deionized water. An initial evaluation of the sorption capability and phosphate release of the original and modified biochars were determined with batch sorption experiments. 0.06 g of each sorbent was added into 30 mL simulated solutions containing 5, 20, and 50 mg L1 phosphate in vitreous vials, and a blank system without phosphate was conducted under the same condition as control. The sorption experiments were performed with three different types of real water samples from nature to identify the selectivity of sorbents for phosphate in practical sorption. Eutrophic river water was sampled from Huzhou (30°390 1900 N 119°370 800 E), and the concentrations were 1.82 mg P L1, 3.53 mg 1 1 NO , 18.83 mg SO2 and 2.64 mg Cl L1. The wastewater 3 L 4 L sample was taken from the effluent of a wastewater treatment plant in Lin’an (30°140 1200 N 119°440 5000 E), and the concentrations 1 were 3.07 mg P L1, 1.46 mg NO , 22.28 mg SO2 L1 and 3 L 4 37.82 mg Cl L1. The piggery wastewater was collected from a pig farm in Hangzhou (30°240 3300 N 119°540 2200 E), and the concen1 trations were 16.71 mg P L1, 0.14 mg NO , 258.26 mg SO2 3 L 4 L1 and 852.18 mg Cl L1. All the vitreous vials were sealed and agitated at 160 rpm on a mechanical shaker for 24 h at 28 ± 0.5 °C to reach apparent equilibrium based on the preliminary study, and the final suspensions were centrifuged and filtered by 0.22 lm filters. Concentrations of phosphate in the filtrates were analyzed by the ascorbic acid method (ESS Method 310.1; (USEPA, 1992)) and the sorbed phosphate amounts were calculated based on their concentration differences between the initial and equilibrium solution. The effect of pH on the sorption of phosphate was investigated with an initial pH range between 2.0 and 10.0, and the initial pH of phosphate solutions was adjusted to the required values by using 0.5 M HCl or NaOH solutions. The final pH and residual phosphate in the solutions were determined after equilibration. Based on the batch sorption experiments, the biochars with higher phosphate sorption capacity were selected for investigating the sorption kinetics and isotherms. Sorption kinetics was evaluated at 28 ± 0.5 °C and the initial pH for each sorption solution was adjusted to 7.0 ± 0.05 by adding 0.5 M HCl or NaOH solutions. The original biochar (0.06 g) was added into 30 mL solutions containing 50 mg L1 phosphate in vitreous vials, and the samples were taken at the interval times after shaken at 160 rpm in a mechanical shaker. Sorption isotherms were determined using batch experiments in vitreous vials under the same conditions mentioned above, and the concentration of phosphate varied from 0 to 100 mg L1. At the end of each experiment, the final suspensions were centrifuged, filtered, and the supernatant solution was separated for analysis of phosphate by the ascorbic acid method

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(ESS Method 310.1; (USEPA, 1992)), and the phosphate loaded biochars were prepared for later analysis. 2.4. Column studies 1.5 cm inner diameter and 30 cm height of PMMA tube was used as a fixed-bed column reactor. The efficiency of modified biochar (TBM-A) bed for phosphate removal was investigated at a certain flow rate (1.5 mL min1) of 20, 50, and 100 mg L1 initial phosphate concentration using 8 cm depth fixed-bed column with approximately 2 g microspheres. Dynamic sorption was conducted by pumping the phosphate solution through the column in a down-flow direction using a peristaltic pump, and 5 mL of sample was collected at a given time interval for analysis. The breakthrough curves were usually expressed by the ratio of outlet and initial concentration of adsorbates (Ct/C0) versus reaction time. 2.5. Statistical analysis The sorption experiments were performed in triplicate, and the standard deviation was obtained by descriptive statistics. The kinetics and isotherms were fitted using Origin Pro 8.0, and R2 values were used to compare the performance of equations. Correlations were analyzed with the Pearson test (two-tailed) at P = 0.01 or 0.05 by SPSS 18.0. 3. Results and discussion 3.1. Physicochemical properties 3.1.1. Fundamental properties All the basic physicochemical properties of the original and modified biochars are presented in Table 1. The yield of raw biochar was 34.61%, which is greater than those derived from the feedstock of common crop residues such as corn stover (17.0%), corn cob (18.9%), orange peel (26.9%), and cottonseed hull (28.9%) produced at the same pyrolysis temperatures (Chen and Chen, 2009; Mullen et al., 2010; Uchimiya et al., 2011), but it slightly decreased after modification with Mg (33.25%). Except for TB-A, the pH values of biochars were around 10 (Table 1), which is consistent with previous reports (Ahmad et al., 2014). The ash in the Mg-modified biochars was higher as compared to the original biochars owing to the formation of Mg-compounds during the pyrolysis. The C, O and H contents of biochars were in the range of 49.56–63.87%, 11.14–26.74% and 2.65–5.21%, respectively. The amendment of Mg, alginate, or chitosan reduced the C content but retained more O, as compared to the original biochar, suggesting that more oxygen-containing functional groups are presented on the modified biochars. The observed atomic ratios of H/C, O/C, and (O + N)/C increased dramatically after modification, especially for biochar microspheres, confirming the introduction of alginate (C6H7NaO6) and chitosan (C6H11NO4). Meanwhile, the increase of O/C in the modified biochars implied that the surface remained more hydrophilic and polar, which is well corresponded to the relative studies (Zhou et al., 2013). As expected, Mg content of the Mg-modified biochars was higher than the original biochars, and the low content of heavy metals indicates that the modified biochars have the potential to be applied as environmental friendly sorbents (Table S1). 3.1.2. Surface morphology The surface morphological structures of the original and modified biochars were shown in Fig. S2. The original biochar (TB) had a smooth surface and very little pores could be observed. After modification with Mg, the fine pore structures were well developed on

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Table 1 Physicochemical properties of the original and modified biochars. Samplea

Yield

pH

(%) TB TBM TB-A TBM-A TB-C TBM-C

34.61 ± 0.37 33.25 ± 0.81 – – – –

9.55 ± 0.01 10.89 ± 0.02 7.95 ± 0.00 10.76 ± 0.01 10.14 ± 0.01 10.42 ± 0.01

SBETb

Ash

Elemental content (%)

Atomic ratio

(%)

N

C

H

O

H/C

O/C

(O + N)/C

(m2 g1)

21.2 ± 0.14 28.3 ± 0.35 17.0 ± 0.00 24.9 ± 0.71 6.0 ± 0.00 13.3 ± 0.35

1.00 0.46 0.85 0.41 4.17 3.54

63.87 54.76 61.89 49.56 58.38 51.49

2.79 2.65 3.62 3.56 5.21 4.93

11.14 13.83 16.64 21.57 26.24 26.74

0.524 0.581 0.702 0.862 1.071 1.149

0.131 0.189 0.202 0.326 0.337 0.389

0.144 0.197 0.213 0.334 0.398 0.448

5.684 116.2 3.995 46.19 2.393 14.72

a TB was the original biochar derived from T. dealbata at 500 °C, and TBM, TB-A, TBM-A, TB-C, and TBM-C were the modified biochars using Mg, alginate, Mg-alginate, chitosan and Mg-chitosan, respectively. b SBET is the Brunauer-Emmett-Teller (BET) surface area, m2 g1.

the surface of TBM, which was consistent with the surface area analysis. The BET specific surface area of TBM was 116.2 m2 g1, which was approximately 20 times greater than that of TB (Table 1). However, after granulation with alginate/chitosan, the surface area of biochar microspheres (2.393–46.19 m2 g1) were lower as compared to original biochar, due either to pore blockage by the alginate/chitosan or partial coating with calcium and sodium on the surface (Roh et al., 2015; Zhou et al., 2013). Correspondingly, it was confirmed by SEM morphology of microspheres, where the surface were covered and encapsulated with adhesive matter (Fig. S2). In addition, the proportion of mesopores (2– 50 nm) varied from 51.6 to 78.2%, implying that mesopores play a dominant role in the pore structure of both original and modified biochars (Table S2). The EDS spectrum demonstrated that the carbon and oxygen dominated the surface of biochars, and magnesium took quite a large proportion in Mg-modified biochars, and these results well coincided with the element analysis (Fig. S2, Tables 1 and S1). 3.1.3. Mineral crystals analysis Several peaks were clearly observed in the XRD spectra of biochars, implying the presence of mineral crystals (Fig. S3). In the spectra of TBM and TBM-A, two strong peaks at 43.0° and 62.3° were identified as magnesium oxide, suggesting that the nanosized crystals on the surface of TBM and TBM-A in the SEM-EDS image were MgO (Fig. S2). While for TBM-C, Mg(OH)2 was the main binding form in the spectrum, and this difference was probably due to the alkaline condition of preparing TBM-C. KCl and NaCl were the typical mineral crystals in the spectrum of the original biochar (TB), which is in high agreement with the EDS spectrum of TB. After modification with alginate, CaCO3 was the dominant mineral, because CaCl2 was introduced as crosslinking agent when preparing TBA microsphere. However, there was minimal amount of crystallites on the surface of TB-C, mainly owing to the encapsulation of chitosan, and this amorphous phase was also the main reason of the lower ash content of TB-C (Table 1). 3.1.4. Functional groups analysis The FTIR spectra of the prepared biochars is presented in Fig. S4. It can be seen that the band around 3404 cm1 represents the stretching vibration of AOH groups, and the peaks at 2921 and 2855 cm1 are assigned to ACH2 groups. The adsorption bands in the wave range of 1620–1300 cm1 correspond to aromatic AC@C, AC@O and ACH2 stretching vibration. The band at 1082 and 1036 cm1 is assigned to CAOAC group, and the peaks of 874, 811 and 749 cm1 can be attributed to the aromatic CAH out-ofplane bending vibrations. The FTIR result clearly showed that some peaks were shifted owing to the introduction of alginate and chitosan (dashed line in Fig. S4), which may affect their sorption capacity (Ahmad et al., 2014). Moreover, more peaks were observed after modification, meaning more oxygen-containing

functional groups formed, in accordance with the results of element analysis (Table 1). Remarkably, agreed with the XRD and SEM-EDS results, the peaks around 3700 and 477 cm1 assigned to AMgAOH and MgAO bonds appeared after modification with Mg, further confirming the presence of Mg in the modified biochars. 3.2. Phosphate sorption 3.2.1. Phosphate release and sorption experiments Phosphate release and sorption capability of the original and modified biochars were determined with batch sorption experiments (Fig. 1). TB showed a greater release capacity for phosphate (0.74 mg g1) than others, and less phosphate was released from TB-A (0.32 mg g1) and TB-C (0.31 mg g1) owing to the immobilization during granulation process. The phosphate release capacity of TBMs (TBM/TBM-A/TBM-C) was almost none (<0.005 mg g1). Obviously, biochar modification with magnesia significantly reduced phosphate release, which was probably benefited from the binding potential of magnesia towards phosphate (Ren et al., 2015a). The results suggested that both granulation with alginate/chitosan and modification with magnesium were helpful in alleviating phosphate release, and biochar modified with Mg had a greater capability of phosphate immobilization. Phosphate sorption capacity of biochars were preliminarily evaluated at the initial concentrations of 5, 20 and 50 mg L1, and the original and modified biochars showed remarkable differences in sorption capacity (Fig. 1b). TB showed no sorption capacity for phosphate, instead it released phosphate even at the high concentration (50 mg L1) in solution, which is consistent with previous studies (Jung et al., 2015; Ren et al., 2015a). Previous studies has reported that most of biochar surfaces are negatively charged, and the low phosphate sorption was driven by electrostatic repulsion interactions between phosphate and alkaline functional groups on biochar surfaces (Uchimiya, 2014). The sorption capacity of TB-A and TB-C was slightly greater than that of TB, indicating that the impregnation of alginate/chitosan may provide more sorbing sites for phosphate, and phosphate may be also trapped in the supporting particle skeleton (Roh et al., 2015; Zhou et al., 2013). By contrast, more than 83.1% of phosphate was removed by TBM from aqueous solution at different phosphate concentrations, and this removal rate increased to 96.8% for TBM-A (Fig. 1b). These results implied that the introduction of magnesia significantly enhanced the sorption capacity of biochar as the introduction of magnesia (pHpzc = 12.0) greatly decreased the negative charge on TB surface (Fig. 2c), thus enhancing phosphate sorption. However, TBM-C showed a relatively lower removal of 13.4–36.8% than TBM and TBM-A. As described above, a lower mineral content and smaller surface area of TBM-C may be responsible for the low phosphate sorption capacity (Yao et al., 2013). Overall, Mg-alginate modified biochar showed the highest phosphate sorp-

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tion capacity, and granulation effectively solved the application and recovery difficulties of biochar resulting from its powder condition. Unlike simulated P-rich solutions, eutrophic water and wastewater from nature are enriched with various ions and organic matter that interfere with phosphate removal (Yin and Kong, 2014). For the eutrophic river water (1.82 mg P L1), all sorbents used in this study showed considerable phosphate removal efficiency (>54.6%), and this removal rate increased to 98.3% and 96.7% for TBM and TBM-A (Fig. 1c). When applied to wastewater samples, TBM and TBM-A exhibited greater phosphate removal (69.0–96.2% and 67.6–94.4%, respectively) than others, but the sorbents without Mg modification had a much lower removal rate (0–11.2%) (Fig. 1c). The results suggested that TBM and TBM-A are effective sorbents for phosphate in the presence of various anions, such as  2 NO 3 , SO4 and Cl . The higher selectivity of these hybrid sorbents for phosphate was probably due to the formation of inner-sphere complexes via Lewis acid-base interactions, while other commonly coexisting anions are only capable of forming outer-sphere complexes through Coulombic interactions (Sarkar et al., 2011). Therefore, Mg-alginate modified biochar evidently has the potential for phosphate removal in water treatment at full scale. 3.2.2. pH dependent sorption of phosphate on biochars As known, pH is considered the major variable that controls phosphate sorption on biochars, as it affects both the molecular form of phosphate in solution and the zeta potential of the biochars (Krishnan and Haridas, 2008). As shown in Fig. 2, no sorption of phosphate occurred at low pH (pH = 2) on TBM and TBM-C, whereas the amount of phosphate sorption on biochars was steady in the initial pH range 3.0–10.0. More than 97.1% of phosphate was removed by TBM-A from aqueous solution with a wide pH range of

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3.0–10.0 for an initial concentration of 50 mg L1, and this removal rate slightly decreased to 88.4% for TBM (Fig. 2), indicating that TBM-A has a wide pH range of application with high phosphate removal rate. Compared with the initial solution, the pH of the equilibrium solution significantly increased (Fig. 2b), especially when the initial pH exceeded 2.0, which could be attributed to the biochar buffering capacity resulting from the protonation/de-protonation of carboxyl and hydroxyl (Uchimiya, 2014). When the initial pH increased from 3.0 to 10.0, the pH of the equilibrium solution was changed in a narrow range from 9.8 to 10.7 due to the superior buffering capacity, and correspondingly the phosphate sorption capacity of TBMs was also steady within this pH range. When solution pH is between 2 and 3.5, H3PO4 and H2PO 4 are the dominant forms in solution, while HPO2 4 is the major species between pH 9 and 11(Krishnan and Haridas, 2008). The electrostatic interaction is regarded as the primary driving force for sorption on biochar surfaces, and the zeta potential of selected biochars are presented in Fig. 2c. It was clearly observed that the introduction of magnesia (pHpzc = 12.0) greatly decreased the negative charge on the TB surface, which enhanced phosphate sorption in the initial solution pH range of 3.0–10.0. However, TBM-C with most positive charge (pHpzc = 8.85) exhibited a lower sorption for phosphate than TBM or TBM-A, indicating that electrostatic interaction is not the dominant mechanism of phosphate sorption on the modified biochars. The lower phosphate sorption capacity of TBMs at pH 2.0 was mainly attributed to the increasing concentration of H3PO4, which is unable to attach to the sites of sorbent. 3.2.3. Phosphate sorption kinetics Kinetics study indicated that the sorption of phosphate on selected biochars increased rapidly in the first 4 h, which

Fig. 1. Phosphate release (a) and sorption (b) capacity of the original and modified biochars from simulated solutions. (c) Phosphate removal capacity of the original and modified biochars from real water samples.

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Fig. 2. (a) Effect of pH on phosphate sorption on selected biochars. (b) The pH values change of the mixed solution of selected biochars with phosphate after equilibration. (c) Zeta potential of the selected biochars at different pH values. Error bars indicated standard deviation.

accounted for around 80% of the equilibrium sorption capacity and reached equilibrium at 24 h (Fig. 3a). Phosphate sorption on TBM occurred more rapidly than TBM-A in the initial 2 h due to the difference in sorbent shape, while TBM-A showed greater equilibrium sorption capacity (23.57 mg g1). The pseudo first order ðQ t ¼ Q e ð1  ek1 t ÞÞ, pseudo second order (Q t ¼ k2 Q 2e t=1 þ k2 Q e t) and intra-particle diffusion (Q t ¼ ki t1=2 þ C) model were applied to describe the sorption data. The pseudo second order model fitted the sorption data on TBM better with the highest squared correlation coefficients (R2 = 0.994), which had a great predicted Qe value (22.22 mg g1) in a good agreement with the experimental data (21.70 mg g1) (Table 2). While for TBM-A, the pseudo first order model described the sorption data better (R2 = 0.990), and the theoretical pseudo first order Qe value (23.79 mg g1) was closer to the experimental data (23.57 mg g1). The fitting result indicated that chemical interactions occurred during the phosphate sorption on TBM, while the phosphate sorption on TBM-A may take place through a physical process (Jung et al., 2015). Unexpectedly, a spot of phosphate was released from TBM-C during the initial sorption process (0–0.5 h), which was replaced by sorption with time prolonged. In addition, intra-particle diffusion model well fitted the data of phosphate sorption on TBM-A and TBM-C within the initial 0.5 h (0.925 < R2 < 0.997) (Table 2, Fig. 3b), suggesting that the intra-particle diffusion was more crucial for phosphate sorption onto biochar microspheres than biochar powder at the early stage.

3.2.4. Phosphate sorption isotherms The Langmuir (Qe = KLQm/(1 + KLCe), Freundlich (Q e ¼ K F C Ne ), and Temkin (Q e ¼ B ln A þ B ln C e ) models were employed to fit the data of phosphate sorption on selected biochars (Fig. 4). Both Langmuir and Freundlich models described the isotherm data well, and the Langmuir model fitted the TBM and TBM-A sorption isotherm better with R2 values of 0.968–0.983 (Table 3). While for TBM-C, the Freundlich model (R2 = 0.985) showed a better fit than the Langmuir model (R2 = 0.973). In general, the Langmuir model better describes sorption on homogeneous surfaces, while the Freundlich model is better for heterogeneous surfaces. Hence, the modeling results implied that alginate modification reduced the homogeneities of biochar surface, but chitosan modification showed opposite change. The maximum sorption capacity (Qm) based on Langmuir model for TBM was 27.63 mg g1 (Table 3), which was larger than those of modified biochars in previous studies (Chen et al., 2011; Park et al., 2015; Ren et al., 2015a), such as ZnCl2 and MgO modified biochars derived from sesame straw (8.67–9.68 mg g1), Fe2O3 modified orange peel derived biochars (0.22–1.24 mg g1), and Fe2O3 loaded biochar derived from cotton stalk (0.96 mg g1). After granulation, the maximum phosphate sorption capacity of TBM-A was almost doubled to 46.56 mg g1. Table 4 presents the comparison of TBM-A with those of other sorbents for phosphate sorption in aqueous solutions, and TBM-A exhibited a greater sorption capacity for phosphate, which could be identified as effective sorbent for

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Fig. 3. (a) Sorption kinetic of phosphate on selected biochars. (b) Solid lines represent the phosphate sorption data fitted by the intra-particle diffusion model in the initial time (0–0.5 h). Symbols are experimental data, and lines are modeled results. Error bars indicate standard deviation.

Table 2 Kinetic parameters of phosphate sorption onto selected biochars obtained from the pseudo-first-order, pseudo-second-order and intra-particle diffusion models. Sample

(mg g TBM TBM-A TBM-C a b c d

Pseudo first orderb

Qexa

21.70 23.57 3.50

1

)

1

K1 (h

)

0.72 ± 0.08 0.30 ± 0.02 –

Qe1 (mg g

Pseudo second orderc 1

)

20.18 ± 0.50 23.79 ± 0.58 –

2

1

R

K2 (g mg

0.980 0.990 –

0.04 ± 0.00 0.01 ± 0.00 –

1

h

)

Qe2 (mg g

Intra-particle diffusion (0–0.5 h)d 1

)

22.22 ± 0.36 28.39 ± 1.49 –

R

2

0.994 0.978 –

Ki (mg g1 h0.5)

C (mg g1)

R2

8.65 ± 1.83 3.45 ± 0.09 2.04 ± 0.28

0.17 ± 0.08 0.21 ± 0.04 1.35 ± 0.13

0.842 0.997 0.925

Qex is the amount of phosphate sorption at equilibrium, mg g1. K1 is the rate constant of pseudo first order model (h1), Qe1 is the sorption capacity at equilibrium calculated by the pseudo first order model, mg g1. K2 is the rate constant of the pseudo second order reaction (g mg1 h1), Qe2 is the sorption capacity at equilibrium calculated by the pseudo second order model, mg g1. Ki is the coefficient of intra-particle diffusion model (mg g1 h0.5).

phosphate. However, the maximum phosphate sorption capacity of TBM-C decreased to 11.53 mg g1 due to a lower mineral content and smaller surface area (Yao et al., 2013). In addition, the RL values for TBM and TBM-A (about 0.01) was much lower than that of TBM-C (0.50), indicating that sorption of phosphate on TBM and TBM-A were more favorable than TBM-C under the experimental conditions. Moreover, the Temkin equation didn’t fit the sorption data well (0.74 < R2 < 0.93), suggesting that phosphate sorption process may be not affected by the sorbate/sorbate interactions. 3.3. Sorption mechanisms 3.3.1. Precipitation with minerals Precipitation occurred between phosphate and dissolved cations such as Ca2+, Mg2+, and Al3+ released from sorbents has been suggested as a major mechanism for phosphate sorption in many studies, especially for the industrial by-product sorbents (steel slag, fly ash, furnace slag, and burnt oil shale) which have a high content of Ca, Mg, Al, and Fe (Lu et al., 2009; Xue et al., 2009). In this study, the pH of mixed solutions changed from 7 to 10 around during the phosphate sorption process, which was favorable for the formation of precipitates. Compared with the original biochar, two different forms of white crystals were clearly observed in the SEM image of P-laden biochars (TBM + P and TBMA + P), and their elemental composition are further investigated by EDS spectrum corresponding the red square region (Fig. S5). The nanosized flakes on the TBM + P surface was identified as Mg-P precipitation by EDS spectrum (Fig. S5), and this precipitate was further determined with typical peaks as Mg3(PO4)2 in the XRD patterns (Fig. S5), suggesting that Mg3(PO4)2 was the major compound of the precipitate on TBM + P surface. Distinguished from TBM + P and other biochar sorbents in previous research, the clusters of granular crystals on TBM-A + P sur-

Fig. 4. Sorption isotherms of phosphate on selected biochars. Symbols are experimental data, and lines are modeled results. Error bars indicate standard deviation.

face was identified as Ca-P precipitates using EDS analysis (Fig. S5). The differences in solution Ca2+ concentration after phosphate sorption confirmed this mechanism. With increasing initial phosphate concentration, the amount of Ca2+ released in the solution were significantly decreased (Fig. S6a), and the correlation between the amount of phosphate sorbed on TBM-A and the amount of Ca2+ decreased in the solution was significant positive (r = 0.955, P < 0.01, Fig. S6b), indicating that Ca-P dominated the phosphate sorption on TBM-A. In addition, the amount of Ca2+ released from TBM-A into the solution was approximately 5–25 times higher than that of TBM and TBM-C mainly due to the intro-

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Table 3 Parameters of phosphate sorption on the selected biochars obtained from the Langmuir, Freundlich and Temkin isotherm models. Sample

Langmuira 1

Qm (mg g TBM TBM-A TBM-C a b c

Freundlichb )

27.63 ± 1.03 46.56 ± 1.83 11.53 ± 2.07

1

KL (L mg

)

0.96 ± 0.19 0.83 ± 0.09 0.01 ± 0.00

2

1

RL

R

KF (mg g

)

0.010 0.012 0.500

0.968 0.983 0.973

10.71 ± 1.13 18.69 ± 1.87 0.36 ± 0.04

Temkinc N

R

B

A (g1)

R2

0.29 ± 0.03 0.35 ± 0.04 0.63 ± 0.04

0.933 0.874 0.985

3.77 ± 0.31 5.17 ± 0.85 0.86 ± 0.12

40.92 ± 6.24 77.20 ± 18.32 2.45 ± 0.72

0.923 0.746 0.814

2

Qm is the maximum sorption capacity (mg g1), KL is the affinity coefficient of Langmuir model, RL indicates the favorable degree of isotherm. KF is the experimentally derived capacity coefficient of Freundlich model (mg g1), N is experimentally derived exponent of Freundlich model. A (g1) and B are the Temkin model constants.

Table 4 Comparison of the Mg-alginate modified biochar with other sorbents for phosphate sorption in aqueous solutions. Adsorbent

Sorption capacity (mg g1)

Reference

Mg-alginate modified biochar Marine macroalgae derived biochar Lanthanum-doped activated carbon fiber Steel slag

46.56 32.58 24.27

Present study Jung et al. (2016) Liu et al. (2011)

16.24

Fe(II)-doped activated carbon Fe(III)-doped activated carbon Aluminum salt slag Fertilizer controlled release agent Electric arc furnace steel slags Basic oxygen furnace steel slags

14.12 8.13 10.63 8.43 0.86 7.63

Xiong et al. (2008) Wang et al. (2011) Ren et al. (2015b) Ma et al. (2011) Barca et al. (2012)

duction of CaCl2 as crosslinking agent, hence resulting in an obvious increment of phosphate sorption by forming Ca-P precipitation. However, Ca-P crystal was not clearly presented in the XRD pattern due to its amorphous structure (Fig. S5). Meanwhile, rare Mg-P precipitation occurred on TBM-A + P, as the solubility product constant (Ksp) of Mg3(PO4)2 was higher than that of Ca3(PO4)2, and thus Ca-P precipitate formed instead.

3.3.2. Ligand exchange Ligand exchange between the unprotonated oxygen atoms of the phosphate and hydroxyls on the surface of biochars is considered as a potential mechanism for phosphate sorption on sorbents (Acelas et al., 2015). TBM-A showed a small phosphate sorption capacity (6.41 mg g1) at pH around 3, and it is unlikely to form above mentioned precipitates at this pH level. TBM-A contains magnesium oxide (Fig. S3), and the AOH ligand may interact with phosphate as reported in previous studies (Acelas et al., 2015; Xue et al., 2009). H2PO 4 is dominant in solution when the solution pH is around 3, hence, the phosphate sorption resulting from ligand exchange could be describe as:

2  Mg  OH þ H2 PO4 () ð MgÞ2 HPO4 þ H2 O þ OH

ð1Þ

Except for the bidentate binuclear complex described in the above equation, the monodentate mononuclear complexes could also be formed with the progress of phosphate sorption (Acelas et al., 2013). In addition, TBM-C had most positive charge (pHpzc = 8.85) but exhibited lower sorption for phosphate than TBM and TBM-A, indicating that electrostatic interaction is not the dominant mechanism of phosphate sorption on the biochars in this study. Therefore, the phosphate sorption on modified biochar microspheres may be more complicated, and the corresponding sorption mechanisms merit further research.

Fig. 5. Breakthrough curves for phosphate sorption onto TMB-A at different initial phosphate concentrations. (Reaction conditions: bed height = 15 cm; flow rate = 2 mL min1).

3.4. Column studies The change of inlet phosphate concentration to the dynamic sorption onto TMB-A was performed at the bed height of 8 cm and flow rate of 1.5 mL min1, as shown in Fig. 5. Initially, the sorption process increased rapidly owing to the availability of active surface sorption sites (Foo and Hameed, 2012). As the phosphate laden solution continued to flow, the accessible sorption sites were gradually occupied, thereby the outlet concentration rose until reaching the saturation points. It was clearly observed that the breakthrough time decreased from 1420 min to 690 min with increasing initial concentration from 20 to 100 mg L1. In addition, with increasing initial concentration, the breakthrough curve became steeper and the length of sorption zone decreased, because of higher phosphate loading rate and diffusion or mass transfer coefficient (Zhang et al., 2011).     t The Yoon-Nelson ln C oCC ¼ K YN t  TK YN , Adam’s-Bohart t     and Thomas models ln CCot ¼ K AB C o t  K AB N o ZF ,     K Th Q Th W Co ln C t  1 ¼ Q  K Th C o t were used to simulate experimental breakthrough curves. Both the Yoon-Nelson and Thomas models fitted the dynamic sorption data better with R2 values of 0.889–0.996 (Table 5), suggesting that both the Yoon-Nelson model which neglects the effect of axial dispersion and the Thomas model which overcomes the limitations of external and internal diffusion are adequate for describing this dynamic sorption process (Foo and Hameed, 2012; Roh et al., 2015). The time required for 50% of phosphate breakthrough (T) obtained from Yoon-Nelson model decreased from 672.3 to 165.9 min with increasing phosphate concentration in the leaching solution, which agrees with

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X. Cui et al. / Bioresource Technology 218 (2016) 1123–1132 Table 5 Parameters of phosphate column sorption on the selected biochars obtained from the Yoon-Nelson, Adam’s-Bohart and Thomas models. C0

Yoon-Nelsona KYN (mL mg

20 50 100 a b c

1

5.29 ± 0.00 5.82 ± 0.22 5.92 ± 0.35

Adam’s-Bohartb min

1

)

2

T (min)

R

672.3 ± 3.34 304.9 ± 6.59 165.9 ± 9.02

0.996 0.954 0.889

1

KAB (mL mg

Thomasc 1

min

0.07 ± 0.00 0.04 ± 0.00 0.01 ± 0.00

)

No (mg L

1

)

2700 ± 41 3986 ± 64 6985 ± 201

R

2

0.873 0.877 0.772

KTh (L mg1 min1)

QTh (mg g1)

R2

0.26 ± 0.00 0.12 ± 0.00 0.06 ± 0.00

10.08 ± 0.05 11.43 ± 0.25 12.44 ± 0.68

0.996 0.954 0.889

KYN is the Yoon-Nelson velocity rate constant and T is the time required for 50% of adsorbate breakthrough. KAB is the adsorption rate constant and N0 is the saturation concentration. KTh is the Thomas rate constant and QTh is the predicted bed capacity.

the experimental result. The bed capacity (QTh) predicted by Thomas model increased with the inlet phosphate concentration, and the QTh values are consistent with the experimental data, especially at the inlet phosphate concentration of 20 mg L1, as evidenced by the higher squared correlation coefficient (R2 = 0.996). However, the Adam’s-Bohart equation didn’t fitted the sorption data well (R2 < 0.9), indicating that the validity of this simple model is limited under the present sorption conditions. 4. Conclusions Mg-alginate modified biochar microspheres derived from Thalia dealbata were successfully fabricated and effectively removed phosphate from aqueous solutions. Compared with the original biochar, physicochemical and surface morphological properties of the modified biochars were significantly changed. Alginate, chitosan and biochar are all low-cost and environmental friendly materials that can be easily obtained from waste materials, so the alginate-biochar microsphere should have economic and environmental advantage over commonly used sorbents. Moreover, after sorption, the P-laden microsphere had a high content of P and K but minimal heavy metals, hence, it has potential as an alternative and renewable fertilizer for crop production. Acknowledgements This study was, in part, supported by a grant from the Ministry of Science and Technology of China (Grant# 2012BAC17B02), the CN-USA International Cooperative Fund (Grant# 2010DFB33960), Ningbo Science and Technology Bureau (Grant# 2012C10003), and Fundamental Research Funds for the Central Universities (Grant# 2014FZA6008). Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.biortech.2016.07. 072. References Acelas, N.Y., Mejia, S.M., Mondragón, F., Flórez, E., 2013. Density functional theory characterization of phosphate and sulfate adsorption on Fe-(hydr)oxide: reactivity, pH effect, estimation of Gibbs free energies, and topological analysis of hydrogen bonds. Comput. Theor. Chem. 1005, 16–24. Acelas, N.Y., Martin, B.D., Lopez, D., Jefferson, B., 2015. Selective removal of phosphate from wastewater using hydrated metal oxides dispersed within anionic exchange media. Chemosphere 119, 1353–1360. Ahmad, M., Rajapaksha, A.U., Lim, J.E., Zhang, M., Bolan, N., Mohan, D., Vithanage, M., Lee, S.S., Ok, Y.S., 2014. Biochar as a sorbent for contaminant management in soil and water: a review. Chemosphere 99, 19–33. Barca, C., Gerente, C., Meyer, D., Chazarenc, F., Andres, Y., 2012. Phosphate removal from synthetic and real wastewater using steel slags produced in Europe. Water Res. 46 (7), 2376–2384.

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