A highly selective polymeric ligand exchanger was developed for the removal of trace As(V) from aqueous solution. This adsorbent was prepared by loading Fe(III) onto porous amidoximated polyacrynitrile (AN)/itaconic acid (IA) copolymers (Fe(III)-AO AN/IA). Negligible ferric ion dissolution was observed from Fe(III)-AO AN/IA in solution of acidic pHs up to 2. As(V) adsorption by Fe(III)-AO AN/IA is a pH-dependent process with maximum capacity of 1.32 mg/g at pH 2–3. The adsorption process was found to be governed by pseudo-second-order kinetics, and could be described by the Freundlich model. Fe(III)-AO AN/IA had higher adsorption selectivity for As(V) than other anions in a simulated groundwater body such as Cl, SO42−, PO43−, SiO32−. Fixed-bed adsorption indicated that As(V) in simulated groundwater could be effectively captured from 400 μg/L to <10 μg/L within 190 bed volumes (BV). The As(V) adsorbed on Fe(III)-AO AN/IA could be efficiently eluted with 10 BV of 5% NaCl solution (at pH = 9.0).

INTRODUCTION

Arsenic contamination has raised great concerns worldwide owing to its toxicity and carcinogenicity (Smith et al. 1992). Long-term exposure to arsenic has been associated with cancer of the skin, lungs, urinary tract, kidneys and liver, and can also produce various other non-cancerous effects (Meng et al. 2000). Therefore, the World Health Organization (WHO) has reduced the guideline for arsenic in drinking water from 50 to 10 μg L−1 and most industrialized countries also take 10 μg L−1 as a statutory limit (Badruzzaman et al. 2004). To secure a safe water environment, various technologies, such as precipitation (Meng et al. 2001), ion exchange (Kim & Benjamin 2004), coagulation–precipitation (Jia & Demopoulos 2008), membranes (Waypa et al. 1997) and adsorption (Yuan et al. 2002; Daus et al. 2004) have been proposed for effective arsenic removal from nature and industrial effluent. Among them, ion exchange is currently an Environmental Protection Agency (EPA)-identified best available technology for removing arsenic because of the high removal efficiency from aqueous solution (EPA 2000). However, the adsorption capacity and selectivity of traditional anion exchange resins for arsenic are severely retarded due to the competition derived from other anions such as sulfate ions, chloride ions, silicate, and phosphate, which are ubiquitously abundant in natural water bodies (Zhao & Sengupta 1998; An et al. 2005). To overcome the drawback of the ion exchange resins, many researchers have devoted their efforts to developing a promising polymeric ligand exchanger (PLE) for the adsorption of arsenic, which is composed of a chelating resin and a transition metal cation (Lewis acid) as its terminal functional group (Henry et al. 2004). PLE has been found to be highly selective for arsenic even in the presence of ubiquitous anions such as sulfate and chloride (Awual et al. 2012). A lot of metal ions, including Fe(III), Cu(II), La(III), Zr(IV) and Al(III), are used to coordinate with various chelating resins for attaining PLE (Matsunaga et al. 1996; Awual et al. 2011; 2012). Among these metal ions, the iron ion is especially favored owing to its widespread, cheapness, and nontoxicity. Thus, a lot of PLE are prepared by loading Fe(III) onto chelating resins for arsenic removal (Matsunaga et al. 1996; Rau et al. 2000; Anirudhan et al. 2011). In general, the properties of the metal-hosting polymeric resin are of critical importance in the preparation of a novel ligand exchanger. The metal-hosting polymeric resins should possess stable physical and chemical properties, and high metal-loading capacity as well as a strong bonding force to metal ions (Henry et al. 2004; An et al. 2010). In our previous paper, porous amidoximated acrylonitrile/itaconic copolymers (AO AN/IA) were prepared and the adsorption properties for Hg(II) were investigated (Ji et al. 2016). The experimental results showed that AO AN/IA exhibited faster adsorption rates for Hg(II) owing to the high hydrophilicity of the polymeric skeleton, which favors the diffusivity and accessibility of Hg(II) to the amidoxime group. Furthermore, it is well known that the amidoxime group is a good ligand and has high capacity for Fe(III) (Lutfor et al. 2000; Dong et al. 2010).

Based on the above considerations, a study of arsenic adsorption using Fe(III)-loaded porous amidoximated acrylonitrile/itaconic copolymers (Fe(III)-AO AN/IA) as adsorbent is presented. The performance of Fe(III)-AO AN/IA for arsenate removal is examined as a function of solution pH, kinetics and thermodynamics. Column adsorption experiments with simulated groundwater are conducted as well to evaluate the practical application potential of Fe(III)-AO AN/IA.

EXPERIMENTAL

Materials and methods

AO AN/IA was synthesized and characterized according to our previous work (Ji et al. 2016). The structure and preparation procedure of AO AN/IA are presented in Figure S1 (available with the online version of this paper).

All other chemicals in this study were reagent grade and used without further purification. The stock solution of 1,000 mg/L As(V) was prepared by dissolving Na2HAsO4·7H2O in deionized water. The pH of the As(V) solution was adjusted by dilute hydrochloric acid or sodium hydroxide.

Porous structure parameters were characterized using an automatic physisorption analyzer ASAP 2020 by Brunauer–Emmett–Teller (BET) and Barrett–Joyner–Halenda (BJH) methods through N2 adsorption at 77 K. Infrared (IR) spectra were recorded on a Nicolet MAGNA-IR550 (series II) spectrometer (Madison, WI); test conditions: potassium bromide pellets, scanning 32 times, resolution 4 cm−1. A pH meter (Mettler-Toledo, LE438 pH, China) was used for the measurement of pH values. The concentration of Fe(III) was measured on a flame atomic absorption spectrophotometer (FAAS) (Model 932A, Australia), equipped with an air-acetylene flame. The concentration of As(V) was determined using a double channel atomic fluorescence spectrometer (AFS-920, China).

Preparation of Fe(III)-AO AN/IA

An amount of 400 mg of AO AN/IA was equilibrated with 400 mL of FeCl3 solution (4 mM) at pH 3 for 24 h. After filtration with filter paper, Fe(III)-AO AN/IA was thoroughly washed with deionized water and dried at 50 °C under vacuum for 48 h. The loaded ferric amount on Fe(III)-AO AN/IA was determined as follows: a given amount of Fe(III)-AO AN/IA was placed into 20 mL of 1M HCl and stirred for 12 h at room temperature. After filtration, the concentration of Fe(III) in the solution was determined by FAAS.

Adsorption procedure

Effect of pH on adsorption

The pH effect on equilibrium uptake of As(V) was examined by adding 50 mg of Fe(III)-AO AN/IA into a flask containing 20 mL of 5 mg/L arsenic solution with varying pH in the range of 1.0–9.0. The mixture was then shaken on a rotating tumbler for 12 h at 25 °C. At equilibrium, the residual concentration of arsenic was determined via AFS. The adsorption capacity was then calculated according to Equation (1): 
formula
1
where Q is the adsorption amount (mg/g), C0 and C are the initial concentration and the concentration of arsenic in solution when the contact time is t respectively (mg/mL), V is the volume (mL), and W is the dry weight of resins (g).

Adsorption kinetics

A kinetic experiment was conducted to test the As(V) adsorption rate and determine the time needed for equilibrium. The experiment was carried out by adding 0.5 g of Fe(III)-AO AN/IA into 200 mL of 5 mg/L As(V) solution. The mixture was shaken continuously at 120 rpm on a rotating tumbler (25 °C). At predetermined time intervals, an aliquot of 1 mL solution was taken and analyzed for As(V). The As(V) uptake at various times was then calculated according to Equation (1).

Adsorption isotherm

The isothermal adsorption was also investigated at different initial arsenic concentrations (0.25–5 mg/L) at pH 6.5 in flasks (25 °C). After a shaking time of 12 h, the residual concentration of arsenic was determined by AFS. The adsorption capacities were calculated also according to Equation (1).

Effect of coexisting anions

The effect of coexisting anions such as chloride, sulfate, silicate and phosphate on the removal of As(V) was investigated by performing the following experiment. An amount of 50 mg of Fe(III)-AO AN/IA was placed to a flask containing 20 mL of 100 μg/L As(V) solution and coexisting ions. The concentration ranges of the coexisting ions were kept at 0–100 mg/L. The pH of the solution was adjusted to 6.5. The mixtures were equilibrated at 25 °C for 12 h with shaking at 120 rpm, and then the concentration of residual As(V) was analyzed by AFS.

Column experiments

The breakthrough behaviors of As(V) as well as various competing anions were tested for Fe(III)-AO AN/IA in a fixed-bed configuration. Experiments were carried out with a glass column (11 mm in diameter and 200 mm in length) equipped with a water bath to maintain a constant temperature. An amount of 1.10 g of Fe(III)-AO AN/IA was packed in the column for all tests. The volume of Fe(III)-AO AN/IA in the resin bed (BV, the total volume of the resin bed) was approximately 5 mL. Synthetic water (simulating real contaminated groundwater) was introduced in the resin bed in a down-flow mode. A peristaltic pump (HL-2, China) was employed to ensure a constant flow rate. The composition of feed water was as follows: As(V) = 400 μg/L, Cl = 25 mg/L, SO42− = 80 mg/L, PO43− = 1 mg/L, SiO32− = 5 mg/L, and pH = 6.5. A constant flow rate of 15 mL/h was maintained. The saturated Fe(III)-AO AN/IA was regenerated by using a binary NaCl-NaOH solution (both 5% in mass).

RESULTS AND DISCUSSION

Characterization of Fe(III)-AO AN/IA

The IR spectra of AO AN/IA and Fe(III)-AO AN/IA are presented in Figure 1. In the spectra of AO AN/IA, the adsorption peaks of 1,646 and 936 cm−1 are assigned to the characteristic peaks of C = N and =N-O-, respectively. Compared with AO AN/IA, the absorption peaks of the C = N and =N-O- groups in Fe(III)-AO AN/IA shift to 1,651 and 939 cm−1, respectively. Furthermore, Fe(III)-AO AN/IA shows a new absorption peak at 1,594 cm−1. The reason for these may be ascribed to the coordination between amidoxime groups and ferric ions. The adsorption isotherms of AO AN/IA were also investigated to evaluate the maximum adsorption capacity of AO AN/IA for Fe(III). The experimental results are shown in Figure 2. The Langmuir (Equation (2)) and Freundlich models (Equation (3)) (Limousin et al. 2007) were used to analyze the equilibrium adsorption isotherm data. The fitted results and corresponding parameters are presented in Figure S2 (available with the online version of this paper) and Table 1. 
formula
2
 
formula
3
where Ce is the equilibrium concentration (mmol/L), qe is the amount adsorbed at equilibrium concentration (mmol/g), q is the maximal adsorption capacity (mmol/g), KL is a binding constant (L/mmol), KF is a constant relating the adsorption capacity (mg L−1), and b is an empirical parameter relating the adsorption intensity, which varies with the heterogeneity of the material.
Table 1

Isotherm parameters for the adsorption of Fe(III) onto AO AN/IA

 Langmuir isotherm model
Freundlich isotherm model
T (°C)q (mmol g−1)KL (L mmol−1)RL2bKF (mmol g−1)RF2
15 1.99 1.18 0.9852 0.40 1.02 0.9800 
25 2.14 1.19 0.9923 0.42 1.23 0.9853 
35 2.63 1.19 0.9982 0.42 1.34 0.9738 
 Langmuir isotherm model
Freundlich isotherm model
T (°C)q (mmol g−1)KL (L mmol−1)RL2bKF (mmol g−1)RF2
15 1.99 1.18 0.9852 0.40 1.02 0.9800 
25 2.14 1.19 0.9923 0.42 1.23 0.9853 
35 2.63 1.19 0.9982 0.42 1.34 0.9738 
Figure 1

The IR spectra of AO AN/IA and Fe(III)-AO AN/IA.

Figure 1

The IR spectra of AO AN/IA and Fe(III)-AO AN/IA.

Figure 2

Adsorption isotherms of Fe(III) on AO AN/IA.

Figure 2

Adsorption isotherms of Fe(III) on AO AN/IA.

From Table 1, it can be noted that the regression coefficient (R2) values of the Langmuir model are higher than the R2 values of the Freundlich model, indicating that the Langmuir model describes the observed data much better than the Freundlich alternative. The maximum adsorption capacities of AO AN/IA for Fe(III) from the Langmuir model at 15, 25 and 35 °C are 1.99, 2.14 and 2.63 mmol/g, respectively. The adsorption capacity obtained from the Langmuir model is consistent with the measured value of the iron-loaded amount of Fe(III)-AO AN/IA (2.12 mmol/g).

The nitrogen adsorption–desorption isotherm results and the salient properties of Fe(III)-AO AN/IA are shown in Figure S3 (available with the online version of this paper) and Table 2, respectively. The adsorption–desorption isotherm in Figure S3 shows that Fe(III)-AO AN/IA is type IV according to the IUPAC classification (Sing et al. 1985), suggesting the presence of mesopores. Compared with AO AN/IA, the surface area of Fe(III)-AO AN/IA decreased from 11.71 to 8.32 m2/g. This can be attributed to the formation of iron oxides in the pores, which blocked the adsorption of nitrogen molecules. In addition, the dissociation of ferric ions from Fe(III)-AO AN/IA was also checked as a function of pH. As seen in Figure 3(a), nearly 8.4% of ferric ions are released from Fe(III)-AO AN/IA at pH 1, while almost no ferric ions are leached from Fe(III)-AO AN/IA in the wide pH range from 1.5 to 12. These results demonstrate that the adsorbent can be used in a wide pH range.
Table 2

Salient properties of AO AN/IA and Fe(III)-AO AN/IA

PropertiesAO AN/IAFe(III)-AO AN/IA
BET surface area (m2 g−111.71 8.32 
Pore volume (cm3 g−10.084 0.021 
Average pore diameter (nm) 120 78 
Fe content (%) 11.84 
Color White Brown 
PropertiesAO AN/IAFe(III)-AO AN/IA
BET surface area (m2 g−111.71 8.32 
Pore volume (cm3 g−10.084 0.021 
Average pore diameter (nm) 120 78 
Fe content (%) 11.84 
Color White Brown 
Figure 3

Effect of pH on (a) dissociation of Fe(III) and (b) adsorption capacities of Fe(III)-AO AN/IA ((As(V) 5 mg/L; Fe(III) AO-AN/IA 50 mg).

Figure 3

Effect of pH on (a) dissociation of Fe(III) and (b) adsorption capacities of Fe(III)-AO AN/IA ((As(V) 5 mg/L; Fe(III) AO-AN/IA 50 mg).

Effect of pH on As(V) adsorption

To clarify the effect of the pH on the uptake of As(V), the adsorption capacities of Fe(III)-AO AN/IA for As(V) were determined in the pH range of 1–9 and the results are depicted in Figure 3(b). Obviously, As(V) is strongly adsorbed onto Fe(III)-AO AN/IA at pH range from 0.5 to 2. The maximum As(V) uptake (1.32 mg/g) is reached at pH 2. In addition, the adsorption capacities of As(V) decrease only gradually as pH increases in the pH range of 3.0–9.0. This phenomenon may be attributed to the competition with OH in aqueous solution at high pH (Tao et al. 2011). On the other hand, lower pH is preferable for As(V) adsorption by ligand exchanger adsorbents, such as Fe(III)-loaded resins (Matsunaga et al. 1996; Gu et al. 2005). At low pH, the main species of arsenic are H3AsO4 and H2AsO4, and the corresponding reaction conducted in As(V) adsorption can be shown as follows (Dzombak & Morel 1987; Muñoz et al. 2002): 
formula

Considering the pH value for most natural wasters falls in the range 6.0–8.0 and that no distinct changes in adsorption capacity of Fe(III)-AO AN/IA for As(V) are observed in the range 2.0–9.0, the subsequent experiments were conducted at pH 6.5.

Adsorption kinetics

The adsorption kinetics of Fe(III)-AO AN/IA for As(V) are presented in Figure 4. The adsorption kinetics data were treated according to the pseudo-second-order model given as Equation (4) (Ho & McKay 1999): 
formula
4
where k2 is the rate constant of pseudo-second-order adsorption (g mg−1 min−1); and qe and qt are the adsorption amounts at equilibrium and at time t, respectively (mg g−1). The correlation coefficient (R2 = 0.9975) demonstrates that arsenic uptake onto Fe(III)-AO AN/IA can be fitted well by the pseudo-second-order kinetic model.
Figure 4

Adsorption kinetics curves of As(V) onto Fe(III)-AO AN/IA (As(V) 5 mg/L; Fe(III) AO-AN/IA 50 mg; pH 6.5).

Figure 4

Adsorption kinetics curves of As(V) onto Fe(III)-AO AN/IA (As(V) 5 mg/L; Fe(III) AO-AN/IA 50 mg; pH 6.5).

As shown in Figure 4, As(V) adsorption exhibits an initially fast step in the first 10 min, which is then followed by a slow process to equilibrium within 150 min. Generally, the adsorption rate found in the literature mainly depends on the characteristics of the hosting chelating resins (Lezzi et al. 1994; Sanchez et al. 2000). Suzuki et al. (1997) synthesized Zr-loaded porous spherical resin and investigated the adsorption properties for As(V). The adsorption equilibrium was attained in 6 h (Suzuki et al. 1997), while for a titanium-dioxide-loaded Amberlite XAD-7 resin 4 h and a Cu(II)-loaded WH-425 resin 7 h were necessary (Tatineni & Hideyuki 2002; Tao et al. 2011). Comparing with the above-mentioned polystyrene backbone resins, the faster adsorption rate of Fe(III)-AO AN/IA for As(V) may be attributed to the higher diffusion rate of the arsenate species to the polymeric ligand in the adsorbent.

Adsorption isotherms

Experimental runs were carried out to assess the effect of As(V) concentration on adsorption performance by Fe(III)-AO AN/IA (25 °C), and the results are illustrated in Figure S4 (available with the online version of this paper). It can be found that the As(V) adsorbed onto Fe(III)-AO AN/IA increases with the increase of the initial As(V) concentration. The Langmuir and Freundlich adsorption isotherm models (Equations (2) and (3)) were employed to fit the adsorption data. The calculated maximum capacity, correlation coefficients, and other isotherm parameters are also shown in Figure S4. These results demonstrate that the regression coefficient (R2) value of the Freundlich model is higher than the R2 value of the Langmuir model, indicating that the Freundlich model describes the observed data much better than the Langmuir alternative.

Competitive adsorption

Generally, some anions such as chloride, sulfate, silicate and phosphate commonly exist in groundwater (Guo & Chen 2005; Manna & Ghosh 2007). These anions might join the competition with As(V) for active sites on the surface of Fe(III)-AO AN/IA. In this paper, the effects of coexisting anions on As(V) removal were assessed by investigating several kinds of anions, such as Cl, SO42−, PO43−, and SiO32− (Hang et al. 2012). The concentration ranges of the selected anions were from 5 to 100 mg/L. As shown in Figure 5, SO42− and Cl ions have little interference in As(V) removal efficiency in the concentration ranges investigated in this study. This phenomenon can be explained as follows: the PLE (Fe(III)-AO AN/IA) was prepared by loading Fe(III) onto AO AN/IA. Since Fe(III) is firmly immobilized on the polymer surface by covalently bonding with amidoxime groups, the positive charges of loaded Fe(III) on the surface of Fe(III)-AO AN/IA remain available to interact with anions in the aqueous phase. Under the experimental conditions (pH 6.5), the predominant species of arsenate, sulfate, silicate, and phosphate are HAsO4, HSO4, H3SiO4, and H2PO4, respectively (Zhao & Stanforth 2001; Genc-Fuhrman & Tjell 2003; Mohan & Pittman 2007). Because of the affinity of HAsO4 as a much stronger ligand than HSO4 and Cl (An et al. 2005), the Lewis-base interactions between As(V) and the immobilized Fe(III) are higher. However, as to H2PO4 and H3SiO4, an obvious decrease of As(V) removal efficiency is observed. The decrease in removal efficiency in the presence of H2PO4 and H3SiO4 is due to the significant competition for adsorption sites with arsenic. This can be explained by the chemical similarity between the two competing ions and arsenic, which leads to significant competition (Pan et al. 2014).
Figure 5

Effect of coexisting anions on the removal of As(V) by Fe(III)-AO AN/IA (As(V) 100 μg/L; coexisting anions 0–100 mg/L; Fe(III)-AO AN/IA 50 mg; pH 6.5).

Figure 5

Effect of coexisting anions on the removal of As(V) by Fe(III)-AO AN/IA (As(V) 100 μg/L; coexisting anions 0–100 mg/L; Fe(III)-AO AN/IA 50 mg; pH 6.5).

Fixed-bed adsorption

In order to evaluate the potential of Fe(III)-AO AN/IA for the treatment of simulated groundwater, the column sorption runs were carried out with a single-component arsenate solution or a multi-component feeding solution through two separate columns filled with 1.1 g of Fe(III)-AO AN/IA, respectively. The breakthrough curves are shown in Figure 6. The breakthrough point is set as 10 μg/L (the horizontal dashed line in Figure 6), which is the maximum contaminant level for arsenic in drinking water promulgated by the World Health Organization (WHO). In the case of simulated groundwater, the effective treatment bed volume (BV) is about 190. Specifically, As(V) in the single-component arsenic solution is efficiently removed by Fe(III)-AO AN/IA within 240 BV. Afterwards, the exhausted Fe(III)-AO AN/IA was subjected to in situ regeneration by using 6% NaCl at pH value of 9.0 (see Figure 7). The As(V) loaded on Fe(III)-AO AN/IA can be completely rinsed within ten bed volumes of regenerate. The column experiment results show that Fe(III)-AO AN/IA is effective for As(V) removal from solution and sorption onto Fe(III)-AO AN/IA may be a potential alternative for the removal of As(V) from water.
Figure 6

Breakthrough curves of As (V) onto Fe(III)-AO AN/IA (pH 6.5; As(V) 400 μg/L; Cl 25 mg/L; SO42− 80 mg/L; PO43− 1 mg/L; SiO32− 5 mg/L).

Figure 6

Breakthrough curves of As (V) onto Fe(III)-AO AN/IA (pH 6.5; As(V) 400 μg/L; Cl 25 mg/L; SO42− 80 mg/L; PO43− 1 mg/L; SiO32− 5 mg/L).

Figure 7

Column dynamic desorption and desorption efficiency curves of As(V) on Fe(III)-AO AN/IA.

Figure 7

Column dynamic desorption and desorption efficiency curves of As(V) on Fe(III)-AO AN/IA.

CONCLUSIONS

An Fe(III)-loaded chelating resin denoted as Fe(III)-AO AN/IA was prepared by immobilizing Fe(III) on amidoximated polyacrynitrile (AN)/itaconic acid (IA) copolymers. The obtained adsorbent was fabricated for arsenate removal from aqueous solution. The loaded iron amount is 2.12 mmol/g. The adsorption capacity of Fe(III)-AO AN/IA for As(V) is 1.32 mg/g. The adsorption kinetics follow a pseudo-second-order rate model. The isotherm adsorption equilibrium can be well described by the Freundlich isotherm model. Fe(III)-AO AN/IA exhibits high selectivity for As(V) even in the presence of high concentrations of anions. Meanwhile, the adsorbed As(V) can be effectively desorbed by a 6% NaCl solution (pH = 9).

ACKNOWLEDGEMENTS

The authors are grateful for the financial support by the Natural Science Foundation of Shandong Province (No. ZR2014EMM016)

REFERENCES

REFERENCES
Awual
M. R.
Shenashen
M. A.
Yaita
T.
Shiwaku
H.
Akinori
J.
2012
Efficient arsenic(V) removal from water by ligand exchange fibrous adsorbent
.
Water Res.
46
,
5541
5550
.
Dzombak
D. A.
Morel
F. M. M.
1987
Adsorption of inorganic pollutants in aquatic systems
.
J. Hydraulic Eng.
113
,
430
475
.
EPA
2000
Arsenic Removal from Drinking Water by Ion Exchange and Activated Alumina Plants (EPA/600/R-00/088)
.
Office of Research and Development
,
Cincinnati, OH, USA
.
Genc-Fuhrman
H.
Tjell
J. C.
2003
Effect of phosphate, silicate, sulfate and bicarbonate on arsenate removal using activated seawater neutralized red mud (Bauxsol)
.
J. Phys. IV
107
,
537
540
.
Ho
Y. S.
McKay
G.
1999
Pseudo-second order model for sorption processes
.
Process Biochem.
34
,
451
465
.
Lezzi
A.
Cobianco
A.
Roggero
A.
1994
Synthesis of thiol chelating resins and their adsorption properties toward heavy metal ions
.
J. Polym. Sci. A Polym. Chem.
32
,
1877
1883
.
Limousin
G.
Gaudet
J. P.
Charlet
L.
Szenknect
S.
Barthès
V.
Krimissa
M.
2007
Sorption isotherms: a review on physical bases, modeling and measurement
.
Appl. Geochem.
22
,
249
275
.
Lutfor
M. R.
Silong
S.
Zin
W. M.
Rahaman
M. Z. A.
Ahmad
M.
Haron
J.
2000
Preparation and characterization of poly(amidoxime) chelating resin from polyacrylonitrile grafted sago starch
.
Eur. Polym. J.
36
,
2105
2113
.
Pan
B. C.
Li
Z. G.
Zhang
Y. Y.
Xu
J. S.
Chen
L.
Dong
H. J.
Zhang
W. M.
2014
Acid and organic resistant nano-hydrated zirconium oxide (HZO)/polystyrene hybrid adsorbent for arsenic removal from water
.
Chem. Eng. J.
248
,
290
296
.
Rau
I.
Gonzalo
A.
Valiente
M.
2000
Arsenate(V) removal from aqueous solution by iron(III) loaded chelating resin
.
J. Radioanal. Nucl. Chem.
246
,
597
600
.
Sing
K. S. W.
Everett
D. H.
Haul
R. A. W.
Moscou
L.
Pierotti
R. A.
Rouquerol
J.
Siemieniewska
T.
1985
Reporting physisorption data for gas/solid systems: with special reference to the determination of surface area and porosity
.
Pure Appl. Chem
.
57
,
603
619
.
Smith
A. H.
Hopenhaynrich
C.
Bates
M. N.
Goeden
H. M.
Hertzpicciotto
I.
Duggan
H. M.
Wood
R.
Kosnett
M. J.
Smith
M. T.
1992
Cancer risks from arsenic in drinking water
.
Environ. Health Perspect.
97
,
259
267
.
Waypa
J. J.
Elimelech
M.
Hering
J. G.
1997
Arsenic removal by RO and NF membranes
.
J. Am. Water Works Assoc.
89
,
102
114
.
Yuan
T.
Hu
J. Y.
Ong
S. L.
Lu
Q. F.
Ng
W. J.
2002
Arsenic removal from household drinking water by adsorption
.
J. Environ. Sci. Health A Tox Hazard. Subst. Environ. Eng
.
37
,
1721
1734
.
Zhao
H. S.
Stanforth
R.
2001
Competitive adsorption of phosphate and arsenate on goethite
.
Environ. Sci. Technol.
35
,
4753
4757
.

Supplementary data