Journal of Hazardous Materials 289 (2015) 54–62

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Preparation and characterization of an organic/inorganic hybrid sorbent (PLE) to enhance selectivity for As(V) Byungryul An a , Hakchan Kim a , Sang-Hyup Lee a,b , Jae-Woo Choi a,c,∗ a b c

Center for Water Resource Cycle Research, Korea Institute of Science and Technology, Hwarangno 14-gil 5 Seongbuk-gu, Seoul 136-791, Republic of Korea Green School, Korea University, 145, Anam-ro, Seongbuk-gu, Seoul 136-701, Republic of Korea Department of Energy and Environmental Engineering, University of Science and Technology (UST), Daejeon 305-350, Republic of Korea

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

• A natural biopolymer, chitosan, was used to selectively remove As(V).

• A deprotonated amino group was coordinated with nickel by LAB.

• Immobilized nickel played a key role in the selective removal of As(V) over SO4 2− . • The breakthrough sequence is followed: As(V) > SO4 2− > HCO3 − > NO3 − > Cl− . • The PLE is reused without a significant loss of As(V) uptake up to 10 times.

a r t i c l e

i n f o

Article history: Received 17 October 2014 Received in revised form 21 January 2015 Accepted 9 February 2015 Available online 11 February 2015 Keywords: Polymeric ligand exchagner Arsenate Selectivity Chitosan

a b s t r a c t For the selective removal of arsenate (As(V)) a hybrid sorbent was prepared using a non-toxic natural organic material, chitosan, by loading a transition metal, nickel. The immobilization of nickel was achieved by coordination with a deprotonated amino group (NH2 ) in the chitosan polymer chain. The amount of nickel was directly correlated to the presence of the amino group and was calculated to be 62 mg/g. FTIR spectra showed a peak shift from 1656 to 1637 cm−1 after Ni2+ loading, indicating the complexation between the amino group and nickel, and a peak of As(V) was observed at 834 cm−1 . An increase of sulfate concentration from 100 mg/L to 200 mg/L did not significantly affect As(V) sorption, and an increase in the concentration of bicarbonate reduced the As(V) uptake by 33%. The optimal pH of the solution was determined at pH 10, which is in accordance with the fraction of HAsO4 2− and AsO4 −3 . According to a fixed column test, a break through behavior of As(V) revealed that selectivity for As(V) was over sulfate. Regeneration using 5% NaCl extended the use of sorbent to up to uses without big loss of sorption capacity. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Arsenic is a contaminant in drinking water of emerging concern because of its serious health effects, such as skin, lung, bladder, and

∗ Corresponding author. Tel.: +82 2 958 5820; fax: +82 2 958 5839. E-mail address: [email protected] (J.-W. Choi). http://dx.doi.org/10.1016/j.jhazmat.2015.02.029 0304-3894/© 2015 Elsevier B.V. All rights reserved.

kidney cancers, even at 5 ␮g/L following long-term exposure [1]. Arsenic is mostly released to groundwater from natural exposure and agricultural applications such as pesticides [2]. 2% of drinking water supplies in USA may have concentrations in excess 20 ␮g/L of arsenic, compared to concentrations in groundwater of 1 ␮g/L [3] and in surface water of less than 1 ␮g/L [2]. In Bangladesh, the presence of inorganic arsenic in drinking water puts 35–77 million people at risk [4], and a British Geological Survey (2001)

B. An et al. / Journal of Hazardous Materials 289 (2015) 54–62

reported that 27% of samples had concentrations over 0.01 mg/L and 46% had over 0.05 mg/L in 61 of the 64 districts [5]. The maximum contaminant level (MCL) in drinking water is set to 10 or 50 ␮g/L, depending on the countries. China, Bangladesh, and Chile have regulated levels of 50 ␮g/L. After the World Health Organization (WHO) recommended a MCL of 10 ␮g/L in 1993 (WHO, 1993) [6], The U.S. (2006) New Zealand (2005), Taiwan (1998), and Canada (2007) complied with this enforced level [7]. The Department of Environmental Protection of New Jersey (USA) adopted the lower current USEPA standard of 5 ␮g/L [8]. Due to the decrease in MCL to protect public health, it is essential to find an optimal method to remove arsenic from drinking water. The US EPA recommended best available technologies (BAT) and an ion exchange (IX) is considered one of effective BATs for sites due to high arsenic removal capacity, easy operation and reusability, its application was only recommended at sulfate levels of less than 50 mg/L, because sulfate is selectively removed over arsenic, which significantly reduces the arsenic removal efficiency [9]. To improve the arsenic selectivity at a high concentration of sulfate, An et al. (2005) [10] used polymeric ligand exchanger (PLE). The concept of PLE was introduced by Helfferich (1962) [11], and PLE has been developed using a commercial chelating resin to remove phosphate [12]. The studies confirmed higher selectivity for arsenic and phosphate even in the presence of high concentrations of sulfate present as a competing ion, as the strongly immobilized transition metal enhanced the arsenic and phosphate removal by Lewis acid and base (LAB). Many studies with a natural polymer as the sorbent have been conducted in water treatment. In particular, alginate and chitosan, which have COOH on C-5 and NH2 on C-5, respectively, were successfully used for removal of heavy metals such as copper [13,14] and nickel [15] with high capacity and an easy synthesis process. Chitosan showed unique properties in water treatment due to the protonation and deprotonation of the amino group depending on the solution pH, which can enable its use for the removal of both cations and anions. The wide range of pKa value from 6.0 to 7.0 is due to the chitosan molecular weight and the degree of deacetylation (DD). Removal of arsenic [16], Cr [17], and dye [18] has been studied using the protonated amino group in chitosan as a means of governing electrostatic interaction, while the deprotonated amino group can interact with heavy metals, especially transition metals, by ligand interaction. Therefore, deprotonated chitosan is regarded as a chelating resin. As a result, chitosan can be an alternative material for PLE in place of a commercial chelating resin. To prepare PLE, Fe3+ was first immobilized to a commercial chelating resin known as Dow 3 N [19] for selective arsenic removal, which did not show high capacity for arsenic removal due to the low Fe3+ loading capacity. Later, Cu2+ was commonly immobilized because of its higher transition metal capacity leading to effective removal of arsenic or phosphate. Instead of a commercial chelating resin, chitosan hydrogel beads can be used as a chelating resin and was used for phosphate removal by loading copper [20,21]. The new approach using nickel (Ni2+ ) which has a higher solubility than copper at all pH range would lead to increase more immobilization than Fe3+ or Cu2+ into chitosan hydrogel which can effectively enhance the arsenic sorption capacity and selectivity, and considering the effect of competing ions can determine the feasibility of chitosan bead. The goal of this research was to prepare highly immobilized chitosan beads and apply them to As(V) removal using a batch and column test. Detail is to (1) prepare PLE based on chitosan with nickel immobilized, (2) characterize the physical and chemical properties through XRD and FTIR spectra, (3) investigate the effects of sulfate and bicarbonate on the As(V) sorption capacity,

55

(4) determine the breakthrough sequence in a multi-component system, and (5) test the reusability of the brine solution. 2. Material and methods 2.1. Chemicals Chitosan flake was purchased from Sigma–Aldrich (Reykjavik, Iceland) with an average molecular weight of 250,000 and a degree of deacetylation of 85%. The solution of arsenate was prepared using NaHAsO4 ·H2 O obtained from Sigma–Aldrich (St. Louis, USA). All other chemicals such as NaHCO3 , NaSO4 , NaNO3 , NaOH, and HCl, are ACS grade (Sigma–Aldrich, USA) and used without further purification. 2.2. Preparation of CB–Ni To synthesize CB–Ni, firstly, chitosan hydrogel beads must be prepared, as well-documented by various studies [20,22]. Nickel was then immobilized into the prepared chitosan hydrogel bead. In brief, 5 g of chitosan flake was added to 200 mL of a 1% (v/v) HCl solution, and the mixture was stirred at 100 rpm for at least 12 h to completely dissolve the chitosan. The chitosan solution was dropped by burette into 1 M of NaOH solution stirred at 100 rpm. The final product was washed several times with DI water until the pH reached a neutral representing CB. The hydrogel beads (∼200 mL) were added to a 5000 mg/L nickel solution with a ratio of 1 to 2 and shaken for 24 h. The formation of the bead is influenced by viscosity, which is a function of the concentrations of chitosan and HCl, and the temperature. 2.3. Physical and chemical characterization of CB–Ni The image of the surface and cross-section, and chemical composition was obtained using scanning electron microscopy (SEM) (InspectTM F50, USA) coupled with energy dispersive X-ray spectrometry (EDX) (HORIBA 6853H, Japan). Before analysis, the chitosan hydrogel beads were first dried in a freezing dryer and coated with platinum using a BAL-TEC SCD005 sputter coater at 60 Ma for 40 s to prevent deformation. Fourier transform infrared spectroscopy (FTIR) spectra (Spotlight 200, PerkinElmer, USA) was used in the range from 380 to 4000 wavenumber (cm−1 ) to evaluate the chemical bonds before and after the arsenic sorption in CB–Ni. The immobilized Ni2+ in hydrogel bead was measured by desorption from CB–Ni using 1% HCl. 2.4. Arsenate sorption equilibrium tests Batch isotherm tests were carried out to find the arsenic uptake for CB–Ni. Each 50 mL of solution containing 10 mg/L of As(V) and 100 or 200 mg/L of sulfate or bicarbonate was prepared at a fixed pH 7 ± 0.2. The desired amount of CB–Ni, ranging from 0 to ∼0.1 g (dried weight), was added to the 50 mL solution. Note that the weight of chiotosan bead was obtained after over dryer at 60 ◦ C after experiment done when all used as gram (g). The mixture was then rotated at 30 rpm for 48 h. To find the As(V) sorption capacity at different pH values, a batch test was performed using 0.05 mg of CB–Ni and 50 mL of solution containing 10 mg/L of As(V) and 100 mg/L of sulfate, with all other conditions maintained the same. The initial pH was adjusted at predetermined time intervals with diluted NaOH or HCl to minimize the shift of pH. 2.5. Fixed bed column test A fixed column test was carried out to determine the breakthrough profiles of As(V) and competing ions such as nitrate,

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B. An et al. / Journal of Hazardous Materials 289 (2015) 54–62

Table 1 The salient physical and chemical properties of CB and CB–Ni. CB

Photo shape Nickel uptake1 (mg/g) Water content (%)2 Swelling rate3 1 2 3

CB2500

CB5000

– –

16 times

– 3 times

Desorbed by 1 M HCl. Calculated by volume. Immersed and shaken in DI water for 10 days.

bicarbonate, and sulfate. An acrylic column (30 mm in diameter and 100 mm in length) was set up with a mode of up-flow. The physical operating parameter, empty bed contact time (EBCT) and flow rate were 18 min and 1 ml/min, respectively. 2.6. Regeneration Batch sorption and desorption series were performed to study the reusability of CB–Ni as a sorbent. All sorption tests were conducted by adding 0.05 (or 0.1 g) of CB–Ni into 50 (100) mL of solution containing 1 mg/L of As(V) and 100 mg/L of sulfate, shaken at 150 rpm for 24 h. After sorption, the As(V) saturated CB–Ni was treated with 5 M NaCl under identical physical conditions. After several washings with DI water, a sorption test was carried out. Due to the hybrid sorbents, the TOC was analyzed to estimate the concentration of chitosan, and the nickel was measured 2.7. Chemical analyses The concentration of arsenate was measured using ICP, which has a detection limit of 10 ␮g/L as arsenic, after being filtered using 0.45 ␮m pore size PTFE membranes (Millipore Corp., USA). An ion chromatograph (ICS-100, USA) was used to measure concentrations of chloride, nitrate, and sulfate. The concentration of bicarbonate (HCO3 − ) was analyzed using a TOC analyzer (TOC-L Hitach, Japan). Solution pH was measured using an Orion star A211 (Thermo Scientific, USA). 3. Results and discussion 3.1. Physical and chemical properties CB and CB–Ni were prepared with physical properties listed in Table 1. After loading Ni2+ into the CB, its white color evenly turned to light green, indicating that the chitosan flake was well dissolved. This means that the amino group was dispersed to prepare the CB and that a molecular of nickel was firmly immobilized in each amino group. The loading amount of Ni2+ in CB–Ni was measured to 62 and 39 mgNi/g-dried chitosan at initial concentrations of 5000 and 2500 mg/L of Ni2+ in solution, respectively. Although there was no proportional relationship of nickel uptake, it is clear that uptake increases with an increase of the loading concentration of nickel. Furthermore, the increased immobilized nickel can be predicted to enhance the arsenate removal from solution, which is evidence that nickel as a functional group in chitosan hydrogel directly plays in role in As(V) sorption. The 62 mg/g nickel uptake is similar to that of copper at the same loading condition. As reported in previous studies [23], chitosan was considered a chelating resin for selective heavy metal removal, and the intrinsic sorption mechanism of ligand exchange is the same of commercial chelating resins. According to the uptake sequence (Cu2+ > Ni2+ ) [24,25], copper is the strongest Lewis acid in a chelating resin, resulting in the highest uptake. How-

ever, in this study, the uptake of Ni2+ and Cu2+ was not influenced by the sequence; this is unique to sorption by chitosan. Unlike commercial chelating resins, chitosan’s sorption capacity and behavior are dependent on the protonation and deprotonation of the amino group in the chitosan polymer chain, so that solution pH is a key parameter in heavy metal sorption. As shown in Eq. (1), transition metals can interact with a deprotonated amino group (N), forming a chitosan-metal complex by LAB interaction. The positive charged amino group (protonation) in Eq. (2) no longer binds with the transition metal; instead, the protonated amino group can interact with anions such as phosphate and dyesas anion IX. Chitosan–NH2 (Deprotonation) + M2+ ↔ Chitosan–NH2 M2+ : possible

(1)

Chitosan–NH3 + (Protonation) + M2+ ↔ Chitosan–NH3 + M2+ : impossible

(2)

Therefore, the sorption capacity for transition metals is strongly affected by the pKa value. Inoue et al. (1993) reported that the maximum capacities of copper and nickel were obtained at pH 4.5 and 5.5, respectively, and the two uptake values were similar [14]. Consequently, the higher nickel solution pH of 5.5 during loading overcomes the Williams and Irvin effects by increasing the fraction of deprotonated amino groups. As reported before [26], chitosan based beads are hydrogels that are cross-linked with water by hydrogen bonding to form three-dimensional polymer networks. To investigate deswelling (dehydration) and the swelling rate of chitosan hydrogel beads, the volume was compared before and after drying by immersion in DI water. Both dried hydrogel beads contained ∼96% water content and were swelled to 3 and 16 times in volume for CB and CB–Ni, respectively. Note that different drying methods, oven drying (at 60 ◦ C) and air drying, showed similar de-swelling and swelling phenomena. The higher swelling rate for CB–Ni is due to the presence of nickel in the polymer chain, which accelerates the hydrogen bond with the water molecules. 3.2. SEM The SEM images show the morphology of the surfaces and crosssections of CB (Fig. 1a and b) and CB–Ni (Fig. 1c and d) taken with 10,000x and 5000x magnifications, respectively. All samples were obtained after freeze drying to avoid the effect of the contraction of the chitosan hydrogel bead. An attempt to increase magnification beyond 10,000x was unsuccessful due to the alteration of the bead as a result of the high voltage. More roughness and wrinkling in the surface were observed for CB–Ni in Fig. 1c, resulting from the interaction between nickel and the amino group. Although microsize (␮m) pores were not clearly shown on the surface, the surface area was reported to be 16 m2 /g [27] and 28.5 m2 /g [20] and high porosity is found in the interior of both chitosan beads, and takes on a hexagonal shape, as shown in Fig. 1b and d. The porosity is the result of phase inversion during liquid–liquid phase separation [28] and has been reported with values ranging from 0.84 [28] to 0.95 [27]. Unlike in Fig. 1b, the result with loading of Ni2+ , shown in Fig. 1d, produces an inner pore size of approximately 2 ␮m with narrow distribution, indicating that the nickel was consistently coordinated with the amino group to form complexation by chelation interaction leading to a rearrangement of the chitosan polymer chain. Energy dispersive X-ray (EDX) and EDX nickel mapping, shown in Fig. 2, provide information on the composition for CB (Fig. 2a) and CB–Ni (Fig. 2b). Over 99 and 87% of element analysis consists of carbon and oxygen for both beads. After loading Ni (II), the peak of nickel appeared and the weight % was obtained to 3.5% for CB–Ni. Nickel was well dispersed, as shown through the EDX nickel mapping, clarifying that CB–Ni’s color was even after being immobilized.

B. An et al. / Journal of Hazardous Materials 289 (2015) 54–62

57

Fig. 1. SEM image of the chitosan hydrogel bead before (left) and after (right) nickel immobilization. Surface (a,c) and cross-section (b,d).

3.3. FTIR In Fig. 3, three FTIR spectra were obtained for CB, bare CB–Ni, and As(V)-laden CB–Ni. The broad and deep peak at 3435 cm−1 for the three samples was assigned to the stretching vibrations of O H from the water and hydroxide groups. For both CB–Ni beads, the weak peak at 3235 cm−1 was attributed to N H stretching vibrations. This peak can sometimes be overlapped by the O H peak (as for CB). Another N H bond clearly appears at 1554 cm−1 for all samples [29]. The peaks at 1159 and 1258 cm−1 represent the C-N stretching from the amino group [30]. A shift from the 1656 cm−1 assigned to O H bond to 1637 cm−1 was observed for CB–Ni and CB–Ni–As(V). A similar shift occurred after copper sorption by chitosan [21]. Therefore, it is assumed that the formation of a complex between nickel and chitosan results in this shift. A new peak at 844 cm−1 for CB–Ni–As(V) is evident upon As(V) sorption. In general, As(V) shows a peak at approximately 800 cm−1 regardless of experimental conditions. Peaks reported for As(V) include 856 cm−1 for Al-oxide [31], 820 cm−1 for Fe–Mn oxide [32], 800 cm−1 for magnetite [34], and 808 cm−1 for titanium dioxide [35]. In addition, Zhang et al. (2005) suggested that As(V) sorption can be a ligand exchanger for Fe–Ce with an increased As(V) peak of M–As–O (M: metal) at 836 cm−1 coupling with weakening of M–OH peak at 1126 cm−1 by decreasing a peak [36]. However, in this study, the As(V) sorption mechanism was not confirmed by FTIR spectra. 3.4. Batch test Batch isotherm tests of As(V) were carried out, as shown in Fig. 4. The observed data for As(V) (symbols) was simulated with the Langmuir (solid line) and Freundlich equations (dot line), and sorption parameters calculated from both equations were listed in Table 2. The Langmuir equation (Eq. 3) and the Frundlich equation (Eq. 4) are shown below, qe =

bQC e 1 + bC e

(3)

where qe and Q are the equilibrium As(V) uptake and the maximum As(V) capacity in the solid phase, respectively, Ce is the equilibrium concentration of As(V) in the aqueous phase (mg/L) and b is the Langmuir affinity coefficient between the sorbent and sorbate. qe = kf C1/n (4) where qe is the arsenate uptake (mg/g) at equilibrium, C is the aqueous arsenate concentration (mg/L) at equilibrium, kf is the adsorption capacity and n measures the sorption intensity. Fig. 4a shows the As(V) isotherm for IRA 900, a commercial ion exchanger, derived from An et al. (2005), who conducted a study

under otherwise identical conditions [10], and for CB and CB–Ni. Maximum uptake values (Q) were obtained of 11.0 and 5.5 mg/g for CB–Ni and IRA 900, respectively. It is very clearly shown that CB–Ni has a much greater capacity than IRA900, confirming that the immobilized nickel in the chitosan chain enhances the removal of As(V) from solution, despite the presence of a high concentration of sulfate. Note that the ignorable uptake of arsenate for CB confirmed that As(V) removal was governed by the presence of Ni(II). The Langmuir equation is better fitted for IRA 900, while for CB–Ni, the Freundlich equation produces a more suitable fit, with an r2 of 0.99. Although the Freundlich equation has some limitations, it demonstrates that As(V) was absorbed at different sites on the complex sorbent of CB–Ni in a process of multi-layer sorption, containing several Langmuir sorption patterns [37]. To investigate the effect of SO4 2− on As(V) sorption uptake, two initial concentrations, 100 mg/L (black circle) and 200 mg/L (open circle) of SO4 2− , were added to 5 mg/L of As(V), as shown in Fig. 4b. As(V) uptake Q values of 11.0 and 11.3 mg/g and kf values of 7.5 and 6.4 were obtained. The doubling of the sulfate concentration had an insignificant effect on As(V) sorption into CB–Ni, in accordance with results obtained from PLEs synthesized based on commercial chelating resin. The negligible sulfate effect has been previously reported by Zhao and SenGupta (1998), who used values of 200 and 400 mg/L of sulfate to study phosphate uptake using PLE prepared with commercial chelating resin [12]. The ability of the deprotonated amino group to coordinate with a transition metal within a chelating resin has been proven, and the firmly loaded transition metal can selectively remove As(V) over sulfate by both LAB and electrostatic interactions. Among competing monovalent anions, bicarbonate is a stronger ligand than either chloride or nitrate [24]. Fig. 4c shows an As(V) isotherm test using initial concentrations of 100 and 200 mg/L of HCO3 − . It may clearly be observed that the higher concentration of HCO3 − decreased the As(V) uptake capacity from 15.2 to 10.4 mg/g. This 33% reduction in the uptake indicates that the sorption of As(V) using CB–Ni is more sensitively to the presence of bicarbonate than to sulfate. This result shows that both As(V) and bicarbonate are stronger ligands than sulfate by LAB. Therefore, as shown in Fig. 5, As(V) and bicarbonate have higher selectivity through LAB interaction force than sulfate, whereas sulfate shows the highest strength by electrostatic interaction. To compare the relation of each contaminant, it is possible to determine the relative affinity of a sorbate in a binary contaminant system using the separation factor (␣) [38]. The arsenate/sulfate (␣As/S ) and arsenate/bicarbonate (␣As/B ) are defined as

˛As/S(B) =

qAs × CS(B) CAs × qS(B)

(5)

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B. An et al. / Journal of Hazardous Materials 289 (2015) 54–62

Fig. 2. SEM–EDX of CB (a) and CB–Ni (b) and EDX mapping of nickel (Ni).

where q and C represent uptake (mg/g) and concentration (mg/L) of each anion in the solid and aqueous phases, respectively. The subscripts of As, S, and B denote arsenate, sulfate, and bicarbonate, respectively. In general, the value of ˛As/S(B) greater than 1 indicates the preference of CB–Ni for As(V) over sulfate or bicarbon-

ate. The given value of the separation factor in Table 2 is greater than unity for all CB–Ni, showing that As(V) is more selective than sulfate or bicarbonate. In contrast, 0.11 for IRA 900 [10] was obtained for commercial IX, which demonstrates that sulfate affinity is greater

B. An et al. / Journal of Hazardous Materials 289 (2015) 54–62

59

CB-Ni-As(V) 1554

Transmittance

844 CB-Ni

3235

1637

3435

CB

1656

4000

3500

3000

2500

2000

1500

1159

1000

500

Wavenumber, cm-1 Fig. 3. FTIR spectra of CB, CB–Ni, and As(V)-loaded CB–Cu.

than that of As(V). As result, the lowest As(V) uptake was due to the interference of sulfate. 3.5. pH effect The solution pH is the most important parameter for the As(V) removal efficiency because it determines the form of the As(V). The pKa values of As(V) are 2.2, 7.0, and 11.5, and the species of As(V), H3 AsO4 , H2 AsO4− , HAsO4 2− , and AsO4 3− , are dependent on the solution pH. For As(V) removal using adsorption on an iron-based sorbent [32,33] and IX [10], the optimized pH values were reported to be acidic and neutral conditions, respectively. It is dependent on the physical and chemical sorption mechanism being overwhelmed. Usually, the use of metal oxide showed that lower pH values caused higher uptakes of arsenate. This is because the positively charged surface at less than pzc in that pH range predominantly leads to As(V) sorption. For the IX process, a pH approximately 7 is optimal for As(V) removal due to the species of As(V). According to Fig. 6, the maximum arsenate efficiency is achieved at a pH of 10 for both amounts of CB-Ni, which is the highest in this research. At 0.04 g of CB–Ni, the effect of pH on As(V) removal was more clearly shown. The optimized pH of 10 and the trend of As(V) removal increasing with an increase in pH have similar trends to the fractions of HAsO4 2− or AsO4 3− drawn in the background in Fig. 6. At a high pH, the increased OH− concentration works as a strongly competing ion. Hence, it appears that the effect of the hydroxide ion was not as significant as in commercial resins. 3.6. Column test Based on the batch isotherm tests and separation factor, it was concluded that arsenic is more favorably absorbed than sulfate and bicarbonate. To confirm the higher selectivity for As(V) than sulfate, Fig. 7 was obtained by a fixed column test carried out to investigate the As(V) breakthrough behavior in the presence of competing ions, such as sulfate, bicarbonate, and nitrate (left yaxis). The concentration of chloride, which is plotted on the right y-axis, immediately reached ∼300 mg/L before sharply falling to zero at 200 BV. Among the anions described, nitrate is the first to reach the initial concentration at 120 BV, followed by bicarbonate and sulfate, which are eluted at 210 and 310 BV, respectively. In comparison, As(V) removal continued even after the competing ions were already saturated. It is noteworthy to evaluate the chromatographic for the competing anion, i.e. the effluent con-

Fig. 4. As(V) sorption isotherms for CB–Ni and commercial IX (IRA 900) (a), at different background concentrations of sulfate (100 mg/L and 200 mg/L) (b) and bicarbonate (c).

centrations of nitrate and sulfate were greater than the influent concentrations, which then maintained the initial concentrations. According to the breakthrough behavior, the affinity for CB–Ni is determined to the following: As(V)(HAsO4 2− , H2 sO4− ) > HCO3− > NO3 − > Cl−

(6)

This sequence of affinity is consistent with the separation factors obtained in Section 3.5. In comparison, Clifford et al. obtained the

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B. An et al. / Journal of Hazardous Materials 289 (2015) 54–62

Table 2 Values of Langmuir (Q, b) and Freundlich (kf , n) parameters, and mean separation factors (SF). Sorbent

Conc.

Langmuir

2−

IRA900 CB–Ni

100 mg/L SO4 100 mg/L SO4 2− 200 mg/L SO4 2− 100 mg/LHCO3 − 200 mg/LHCO3 −

Freundlich

SF

Q

b

r2

kf

n

r2

␣As/S(B)

5.5(0.24) 11.0(0.68) 11.3(0.58) 15.2(1.1) 10.4(0.45)

2.4(0.34) 2.9(0.67) 1.4(0.23) 5.3(1.5) 3.0(0.55)

0.98 0.93 0.98 0.90 0.96

3.5(0.22) 7.5(0.084) 6.4(0.062) 11.4(0.13) 7.2(0.047)

3.4(0.68) 3.5(0.14) 3.1(0.10) 3.6(0.11) 4.0(0.10)

0.93 0.99 0.99 0.99 0.99

0.20(0.057) 6.9(0.80) 7.1(2.2) 20.6(4.8) 11.5(2.2)

100

100

80

80

HAsO4

60

H2AsO4

2-

-

0.6g 0.4g CB

60

40

40

20

20

0

illustrated at neutral solution (pH 7)

As(V) species fraction, %

Removal efficiency As(V), %

Fig. 5. A schematic illustration of the selective binding of anions onto CB–Ni.

0 5

6

7

8

9

10

pH Fig. 6. The effect of pH on As(V) uptake (solid) and the fraction of As(V) (dot).

following sequence of affinity for typical strong anion IX [39]: SO4 2− > HAsO4 2− > CO3 2− , NO3 − > Cl− > H2 AsO4 − , HCO3 −  H3 AsO

(7)

The sequence for IX is clearly different than that of with PLE. The sulfate showed the highest affinity of any anion, while nitrate and chloride demonstrated higher selectivity than bicarbonate. The selectivity of anionic contaminants was determined by the driving force that is applied between the sorbate and the sorbent. Consequently, LAB interactions dominates ligand exchange for As(V) or bicarbonate, while the ion exchange was governed by electrostatic interactions. 3.7. Regeneration Batch sorption and desorption of As(V) was performed to investigate sorbent reusability for cost-effectiveness and environmental friendliness. NaCl is a highly effective regenerant for most anion saturated ion exchangers [39]. It was also used in PLE regeneration for As(V) [10] and phosphate desorption [12]. The As(V)-saturated CB–Ni was regenerated using 5% NaCl. The regeneration reaction is

2+ − − + C B − N i2+ (H A sO2− 4 )or C B − N i (H2 A sO4 )2 + 2Cl + xH 2− − → C B − N i2+ C l− 2 + (1 − x)H A sO4 + xH2 AsO4

(8)

At ∼pH 7, the species of As(V) mainly exists with HAsO4 2− or H2 AsO4 − because the pKa2 is 7.2 (Fig. 6). Fig. 8 shows the As(V) uptake capacity (mg/g) after 10 cycles of regeneration compared to that of fresh PLE (left y-axis), with the concentration of TOC in solution described on the right y-axis. The arsenic uptake decreased by ∼15% after two regenerations and held steady within 80% of the initial sorption capacity until 10 cycles. Reduction of sorption efficiency generally manifests by increasing regeneration times. Reduction of sorption efficiency is caused by the loss of unstably immobilized nickel from the chitosan beads which was measured by 10.4 and 1.7% at first and second regeneration, respectively. Extremely highly concentrated sodium chloride replaces As(V) within the saturated CB–Ni and washes some nickel molecular from the amino group. As a result, the reduced nickel affects the reduction in capacity for As(V). As a natural polymer applied in the field of water treatment, persistence of physical and chemical properties is a concern, despite its non-toxicity. In Fig. 8, the range of TOC as an indicator of chitosan amount dissolved ranges from 1 to 1.5 mg/L, regardless of cycle number. Consequently, the destruction of the chitosan bead can be ignored and CB–Ni can be considered a reusable sorbent for As(V) removal with high regeneration efficiency. 4. Conclusions A selective sorbent for As(V) was successfully synthesized on chitosan hydrogel beads prepared by phase inversion using immobilizing nickel as a functional group. The use of CB–Ni enhanced the affinity of As(V) over sulfate, as confirmed through batch and column tests. The major findings and conclusions are summarized as follows: During the immobilization of the transition metal, the capacity was governed by Erving order and the solution pH, which determines the fraction of protonated (NH3 + ) and deprotonated

B. An et al. / Journal of Hazardous Materials 289 (2015) 54–62

350

2.0

1.5

250

NO3-

200 1.0 150 SO42-

100

0.5

HCO3

As(V)

50 0.0 0

200

400

600

800

1000

Concentratino of Cl-, mg/L

300

Cl-

C/C0

61

0 1200

BV Fig. 7. Breakthrough behavior of As(V) and competing ions within a fixed-bed column using a CB–Cu. 2.0

Acknowledgments

As(V) uptake, mg/g TOC, mg/L

1.5

8

6

1.0

4 0.5 2

0

TOC, mg/L in solution

As(V) uptake, mg/g

10

0.0 fresh 1

2

3

4

5

6

7

8

9

10

Regeneration frequency Fig. 8. As(V) removal capacity (left y-axis) and concentration of TOC (right y-axis) using 5% NaCl regenerant for 10 cycles of regeneration.

(NH2 ) amino groups on the chitosan polymer. Deprotonated amino groups can interact with transition metals only by LAB. A maximum As(V) sorption capacity of 11.0 mg/g was obtained for CB–Ni, which is two times higher than that of IRA 900, commercial IX. The increased capacity for As(V) is because it can be removed by both LAB and electrostatic interaction. The effect of the increase of sulfate concentration from 100 to 200 mg/L has no impact on the As(V) sorption capacity, but the doubling of the bicarbonate reduces the As(V) sorption capacity by 33%, as bicarbonate can competes with As(V) as a ligand. The solution pH governed the form of As(V), and the As(V) removal efficiency was increased as the pH increased. The optimal pH was determined to be 10, and the removal efficiency was exactly parallel to the fractions of HAsO4 2− and AsO4 3− . From the fixed column test, the breakthrough sequence was obtained: As(V) > sulfate > bicarbonate > nitrate > chloride, which proved that the affinity of As(V) is greater than that of sulfate and bicarbonate, which were present in concentrations 100 times higher. As(V)-laden CB–Ni was efficiently regenerated by 5% NaCl. The resorption capacity was recovered to 90%, and it persisted after 10 cycles.

This work was supported by Nano-Convergence Foundation (www.nanotech2020.org) funded by the Ministry of Science, ICT and Future Planning (MSIP, Korea) & the Ministry of Trade, Industry and Energy (MOTIE, Korea) [Project Number: R201400210] and the Korea Institute of Science and Technology(KIST) institutional program (2E25311). References [1] NRC, Arsenic in drinking water: Updata, in, National academy press, Washington D.C, 2001. [2] A.H. Welch, D.B. Westjohn, D.R. Helsel, R.B. Wanty, Arsenic in ground water of the United States: Occurrence and geochemistry, Ground Water 38 (2000) 589–604. [3] ATSDR, Public health statement for arsenic, In: US DOHAHS (Ed.), Atlanta, GA, 2007. [4] A.H. Smith, E.O. Lingas, M. Rahman, Contamination of drinking-water by arsenic in Bangladesh: a public health emergency, B World Health Organ 78 (2000) 1093–1103. [5] BGS, Arsenic contamination of groundwater in bangladesh, in: B.G. Survey (Ed.), BGS Technical Report WC/00/19, Keyworth, UK, 2001. [6] WHO, Guidelines for drinking-water quality, 2nd ed., World Health Organization, Geneva, Swichland, 1993. [7] K. Henke, Arsenic: Environmental Chemistry, Health Threats and Waste Treatment, Wiley, 2009, 2015. [8] NJDEP, NJGS Information circular: Arsenic in new jersey ground water, In: D.O.E. protection (Ed.), State of New Jersey, New Jersey, 2004. [9] USEPA, National Primary Drinking Water Regulations; Arsenic and clarifications to compliance and new source contaminants monitoring, In, Environmental Protection Agency, 2001, pp. 6976-7066. [10] B. An, T.R. Steinwinder, D.Y. Zhao, Selective removal of arsenate from drinking water using a polymeric ligand exchanger, Water Res. 39 (2005) 4993–5004. [11] F. Helfferich, Ligand exchange. I. Equilibria, J. Am. Chem. Soc. 84 (1962) 3237–3242. [12] D.Y. Zhao, A.K. Sengupta, Ultimate removal of phosphate from wastewater using a new class of polymeric ion exchangers, Water Res. 32 (1998) 1613–1625. [13] B. An, H. Son, J. Chung, J.W. Choi, S.H. Lee, S.W. Hong, Calcium and hydrogen effects during sorption of copper onto an alginate-based ion exchanger: Batch and fixed-bed column studies, Chem. Eng. J. 232 (2013) 51–58. [14] K. Inoue, Y. Baba, K. Yoshizuka, Adsorption of metal-Ions on chitosan and cross-linked copper(II)-complexed chitosan, Bull. Chem. Soc. Jpn. 66 (1993) 2915–2921. [15] Y. Vijaya, S.R. Popuri, V.M. Boddu, A. Krishnaiah, Modified chitosan and calcium alginate biopolymer sorbents for removal of nickel (II) through adsorption, Carbohydr. Polym. 72 (2008) 261–271. [16] C.C. Chen, Y.C. Chung, Arsenic removal using a biopolymer chitosan sorbent, J. Environ. Sci. Health A. 41 (2006) 645–658. [17] P. Baroni, R.S. Vieira, E. Meneghetti, M.G.C. da Silva, M.M. Beppu, Evaluation of batch adsorption of chromium ions on natural and crosslinked chitosan membranes, J. Hazard. Mater. 152 (2008) 1155–1163.

62

B. An et al. / Journal of Hazardous Materials 289 (2015) 54–62

[18] D. Xu, S. Hein, S.L. Loo, K. Wang, The fixed-bed study of dye removal on chitosan beads at high pH, Ind. Eng. Chem. Res. 47 (2008) 8796–8800. [19] M. Chanda, K.F. Odriscoll, G.L. Rempel, Ligand-exchange sorption of arsenate and arsenite anions by chelating resins in ferric ion form.2. Iminodiacetic chelating resin chelex-100, React. Polym. 8 (1988) 85–95. [20] B. An, K.-Y. Jung, S.-H. Lee, S. Lee, J.-W. Choi, Effective phospahte removal from synthesized wastewater using copper–chitosan bead:Batch and fixed-bed column studies, Water Air Soil Pollut. 225 (2014) 2050. [21] J. Dai, H. Yang, H. Yan, Y.G. Shangguan, Q.A. Zheng, R.S. Cheng, Phosphate adsorption from aqueous solutions by disused adsorbents: chitosan hydrogel beads after the removal of copper(II), Chem. Eng. J. 166 (2011) 970–977. [22] E. Guibal, C. Milot, J.M. Tobin, Metal-anion sorption by chitosan beads: Equilibrium and kinetic studies, Ind. Eng. Chem. Res. 37 (1998) 1454–1463. [23] Y. Kawamura, M. Mitsuhashi, H. Tanibe, H. Yoshida, Adsorption of metal-Ions on polyaminated highly porous chitosan chelating resin, Ind. Eng. Chem. Res. 32 (1993) 386–391. [24] D.Y. Zhao, A.K. SenGupta, Ligand separation with a copper(II)-loaded polymeric ligand exchanger, Ind. Eng. Chem. Res. 39 (2000) 455–462. [25] H. Williams, R.J.P. Irving, The stability of transition-metal complexes, J. Chem. Soc. (1953) 3192–3210. [26] B.L. Guo, Q.Y. Gao, Preparation and properties of a pH/temperature-responsive carboxymethyl chitosan/poly(N-isopropylacrylamide) semi-IPN hydrogel for oral delivery of drugs, Carbohydr. Res. 342 (2007) 2416–2422. [27] Z. Modrzejewska, Sorption mechanism of copper in chitosan hydrogel, React. Funct. Polym. 73 (2013) 719–729. [28] F. Zhao, B.Y. Yu, Z.G. Yue, T. Wang, M. Wen, Z.B. Liu, C.S. Zhao, Preparation of porous chitosan gel beads for copper(II) ion adsorption, J. Hazard. Mater. 147 (2007) 67–73. [29] A. Pawlak, A. Mucha, Thermogravimetric and FTIR studies of chitosan blends, Thermochim. Acta 396 (2003) 153–166.

[30] A. Khan, M.B.H. Othman, K.A. Razak, H.M. Akil, Synthesis and physicochemical investigation of chitosan-PMAA-based dual-responsive hydrogels, J. Polym. Res. 20 (2013). [31] S. Goldberg, C.T. Johnston, Mechanisms of arsenic adsorption on amorphous oxides evaluated using macroscopic measurements, vibrational spectroscopy, and surface complexation modeling, J. Colloid. Interface Sci. 234 (2001) 204–216. [32] B. An, D.Y. Zhao, Immobilization of As(III) in soil and groundwater using a new class of polysaccharide stabilized Fe-Mn oxide nanoparticles, J. Hazard. Mater. 211 (2012) 332–341. [33] G. Sheng, H. Dong, R. Shen, Y. Li, Microscopic insights into the temperature-dependent adsorption of Eu(III) onto titanate nanotubes studied by FTIR, XPS, XAFS and batch technique, Chem. Eng. J. 217 (2013) 486–494. [34] B. An, Q.Q. Liang, D.Y. Zhao, Removal of arsenic(V) from spent ion exchange brine using a new class of starch-bridged magnetite nanoparticles, Water Res. 45 (2011) 1961–1972. [35] M. Pena, X.G. Meng, G.P. Korfiatis, C.Y. Jing, Adsorption mechanism of arsenic on nanocrystalline titanium dioxide, Environ. Sci. Technol. 40 (2006) 1257–1262. [36] Y. Zhang, M. Yang, X.M. Dou, H. He, D.S. Wang, Arsenate adsorption on an Fe–Ce bimetal oxide adsorbent: Role of surface properties, Environ. Sci. Technol. 39 (2005) 7246–7253. [37] A. Delle Site, Factors affecting sorption of organic compounds in natural sorbent/water systems and sorption coefficients for selected pollutants. A review, J. Phys. Chem. Ref. Data 30 (2001) 187–439. [38] F. Helfferich, Ion exchagne, Dover Publications, Inc-New York, 1962. [39] D.A. Clifford, G. Ghurye, A.R. Tripp, Development of an anion exchange process for arsenic removal from water, Arsenic Exposure and Health Effects (1999) 379–387.

inorganic hybrid sorbent (PLE) to enhance selectivity for As(V).

For the selective removal of arsenate (As(V)) a hybrid sorbent was prepared using a non-toxic natural organic material, chitosan, by loading a transit...
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