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Enhanced arsenite removal from water by radially porous Fe-chitosan beads: Adsorption and H2O2 catalytic oxidation
Journal of Hazardous Materials 373 (2019) 97–105
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Enhanced arsenite removal from water by radially porous Fe-chitosan beads: Adsorption and H2O2 catalytic oxidation
T
Yuanfeng Weia, Xingwen Yua, Chengbin Liua, , Jianhong Mab, Shudan Weia, Tao Chena, Kai Yina, Hui Liua, Shenglian Luoa ⁎
a b
State Key Laboratory of Chemo/Biosensing and Chemometrics, Hunan University, Changsha, 410082, PR China College of Environmental Science and Engineering, Hunan University, Changsha, 410082, PR China
GRAPHICAL ABSTRACT
ARTICLE INFO
ABSTRACT
Keywords: Fe-chitosan bead H2O2 Arsenite Adsorption Fix-bed process
Although Fe-chitosan adsorbents are attractive for removing arsenite from water, the practical applications of these granular adsorbents are mainly limited by slow adsorption kinetics. In this study, radially porous Fechitosan beads (P/Fe-CB) were prepared using freeze-casting technique. The P/Fe-CB particles possess radially aligned micron-sized tunnels from the surface to the inside as well as excellent acid resistance. Kinetic studies show that the adsorption equilibrium time of P/Fe-CB to 0.975 mg/L As(III) (within 240 min) is considerably shorter than that of compact Fe-chitosan beads (over 600 min). The maximal adsorption capacity of P/Fe-CB for As(III) is 52.7 mg/g. It can work effectively in a wide pH range from 3 to 9, and the coexisting sulfate, carbonate, silicate and humic acid have no significant effect on As(III) removal. The addition of H2O2 can further accelerate and promote the As(III) removal except at high pH (11) and phosphate concentration (50 mg/L). The fixed-bed experiments demonstrate that the P/Fe-CB column can effectively treat about 3000 bed volume (BV) of simulated As(III)-containing groundwater to meet the drinking water standard (< 10 μg As/L). This study would extend the potential applicability of porous Fe based chitosan adsorbent and millimeter-sized adsorbent combined with H2O2 to a great extent.
⁎
Corresponding author. E-mail address: [email protected] (C. Liu).
https://doi.org/10.1016/j.jhazmat.2019.03.070 Received 18 December 2018; Received in revised form 26 February 2019; Accepted 16 March 2019 Available online 18 March 2019 0304-3894/ © 2019 Elsevier B.V. All rights reserved.
Journal of Hazardous Materials 373 (2019) 97–105
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1. Introduction
with H2O2 for As(III) removal.
Arsenic as a carcinogenic element has become a global threat to the surface water and groundwater due to the natural and anthropogenic release [1]. Arsenic contaminated water is served as the primary drinking water source in many regions [2,3]. To minimize the risk of arsenic, the World Health Organization (WHO) lowered the maximum contaminant level (MCL) of arsenic in drinking water to 10 μg L−1. Arsenic exists mostly in two inorganic forms as oxyanions of arsenite As (III) and arsenate As(V) in natural groundwater. The dominant arsenic species in groundwater is As(III), which is 25˜60 times more toxic than As(V) [4]. Hence, removal of As(III) from the affected water is crucial. Iron (oxyhydr)oxides have been extensively used for arsenic removal, owing to their promising properties such as high sorption affinity toward arsenic, low cost, rich source, and environmental benignity [5,6]. Iron (oxyhydr)oxides usually appear as fine or ultrafine powders which are difficult to apply to fixed-bed columns due to the excessive pressure drop. To achieve a wider range of applications, iron (oxyhydr) oxides are usually immobilized on a support to eliminate this defect [7–9]. Fe-chitosan beads for arsenic removal hold considerable promise due to its practical advantages [10–12]. Chitosan is an abundant waste product that is derived from insect and shellfish exoskeletons [13]. It possesses many advantages such as inexpensive, biodegradable, and non-toxic [14]. Moreover, it can be easily formulated into beads. Despite extension in iron (oxyhydr)oxides application, Fe-chitosan beads usually shown slow kinetics due to its low porosity. While, practical column filtration typically has a short empty bed contact time (EBCT) [15]. Creation of porous chitosan beads is a workable solution to speed up the internal diffusion, thereby speeding up reaction kinetics. Moreover, this operation makes more adsorption sites accessible and thus improvement in adsorption capacity. Freeze-casting is an effortless technique of making porous structure, which exploits endogenous ice crystals as templates to shape and press building blocks to achieve a porous structure [16–18]. This technique does not require additional reagents. Moreover, the resulted pores are aligned [19], which can reduce the ionic diffusion path, facilitate ionic motion to the inner part of the bead, and improve utilization of adsorption sites. In addition to improving the adsorbent, oxidation of uncharged As (III) to As(V) anion is an important strategy to accelerate the As(III) removal as well as increase the As(III) adsorption capacity, because As (V) has a stronger affinity to most of adsorbents relative to As(III) [6,20–24]. Among various oxidants, hydrogen peroxide (H2O2) is the most benign oxidant because its end product is H2O. However, the oxidation activity of H2O2 is strongly pH-dependent and limited to alkaline condition [25]. Integration of H2O2 with iron oxides can generate powerful reactive species even under acidic conditions [26,27]. Despite all this, only few studies focus on metal oxide/H2O2 for As(III) removal [28–30]. Additionally, these studies are limited to nanosized metal oxide and batch experiments. Furthermore, the comparison of As (III) removal by metal oxide with or without H2O2 has not been systematically investigated in the presence of coexisting ions and at different pH conditions. Herein, we fabricated a porous Fe-chitosan bead (P/Fe-CB) using a simple freeze-casting technique. Additionally, the P/Fe-CB was chemically cross-linked by epichlorhydrine (ECH) to markedly improve its acid resistance, which is usually ignored for Fe-based adsorbents. This P/Fe-CB possesses some advantages: fast adsorption, high adsorption capacity and excellent acid resistance. A heterogeneous catalytic oxidation process was also developed by integrating P/Fe-CB and H2O2 to further promote and accelerate As(III) removal. The effect of H2O2 on As(III) removal by P/Fe-CB was systematically investigated. Both batch and column experiments were carried out to examine its adsorption characteristics and to evaluate the potential for practical application. This work would open up a new avenue for the construction of porous Fe based adsorbents and the application of millimeter-sized adsorbents
2. Materials and methods 2.1. Chemicals All chemicals were analytical grade. Chitosan powder (deacetylation: 95%) was purchased from Aladdin Chemistry Co., Ltd. Ferric chloride (FeCl3⋅6H2O) and epichlorohydrin (ECH) were purchased from Sinopharm Chemical Reagent Co., Ltd., Shanghai, China. As(III) stock solution was prepared with NaOH solution using As2O3. As(III) working solution was freshly prepared by diluting arsenic stock solutions with deionized water. 2.2. Preparation of porous Fe-chitosan beads 1 g of chitosan was completely dissolved in 30 mL of acetic acid solution (2% (v/v)). Then, 4 mL of 0.46, 0.92, 1.85 or 2.77 M FeCl3 solution was added into the solution and stirred until it became homogenous. 1 mL of ECH was then added into the homogenous solution under vigorous agitation. The resulting solution was pushed through a 10-mL syringe fitted with 8 G needles into 0.5 M NaOH solution to form beads. The resultant beads were kept for 24 h in the NaOH bath and then washed with deionized water until the residual water was neutral. The wet beads were placed in an ultra-low temperature freezer (−71 °C) for 1 d. After that, these beads dried by freeze-drying. The beads are abbreviated as P/Fe-CB-X (X = 0.5, 1, 2, 3), where X stands for the mass ratio of FeCl3 / chitosan. For comparison, the samples naturally dried at room temperature was prepared and abbreviated as Fe-CB-X (X = 0.5, 1, 2, 3). 2.3. Characterizations The morphologies of the samples were observed on a Hitachi S-4800 scanning electron microscope operating at 3.0 kV. The crystal phases of the samples were collected on an X-ray diffractometer with Cu-Kα radiation (XRD, M21X, MAC Science Ltd.). The surface chemistry properties of the adsorbent before and after the adsorption process were determined by X-ray photoelectron spectroscopy (XPS, K-α 1063, Thermo Fisher Scientific) and Fourier transform infrared spectroscopy (FTIR, Bruker TENSOR 27). The pH values at the point of zero charge (pHPZC) of the adsorbent were measured in a series of 0.01 M NaCl with different initial pH. The Brunner − Emmet − Teller (BET) specific surface areas were measured on a Belsorp-Mini II analyser. 2.4. Batch adsorption experiments Most experiments were carried out in a 50-mL glass bottle containing 30 mL of 0.975 mg/L arsenite solution. During the adsorption process, the reaction system was shaken in a thermostatic orbit shaker at 180 rpm under 25 °C. The adsorbent dose was 1 g/L and the initial pH values of the arsenic solutions were adjusted with HCl solution or NaOH solution. The ionic strength in the solution was adjusted by the addition of NaCl. The effect of humic acid and ions was examined by adding preset amount of humic acid, Na2CO3, Na2SO4, Na2SiO3 and Na3PO4 into the test solution. To make the pH constant during the reaction, moderate amount of NaOH or HCl was added every three hours. The concentration of residual arsenic was determined by an atomic fluorescence spectrophotometer (AFS-9700, Beijing Haiguang Instrument Inc., China). 2.5. Column experiments Column adsorption experiments were carried out at ambient temperature using P/Fe-CB (4.3 g) packed in fixed-bed columns (12 mm in diameter and 100 mm in length). The schematic flowchart is presented 98
Journal of Hazardous Materials 373 (2019) 97–105
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in Fig.S1 (Supporting Information). About 10 mm thick of silica wool was used at each end of the column to distribute the flow. The effective bed volume (BV) of the column was 9.04 mL A peristaltic pump was used to maintain an upward flow. The empty bed contact time (EBCT) was set as 4.5 min and the corresponded flow rate was 2.0 mL/min.
CB-1.0, P/Fe-CB-2.0 and P/Fe-CB-3.0 samples possessed similar porous structures (Fig. S3). The characteristic porous structures would substantially accelerate the transfer rate of arsenite inside the adsorbent. In contrast, the naturally dried Fe-CB-0.5 exhibited a quite different surface and internal structure (Fig. 1d-f). The surface of Fe-CB-0.5 beads was rough and its inside was compact. The BET surface areas of all the sample are summarized in Table S1. The BET surface area of P/Fe-CB beads was much larger than that of Fe-CB beads. The XRD pattern of P/ Fe-CB-0.5 is shown in Fig. S4. The broad diffraction peak at 20.2° is ascribed to chitosan [31], while the peaks at approximately 35.1 and 62.2° correspond to the diffraction of the (110) and (115) planes of ordered 2-line ferrihydrite [32]. Chitosan and ferrihydrite (iron oxide) are readily soluble in dilute acidic solutions, while some of the As-contaminated wastewater are acidic, such as the acid mine drainage (AMD). Therefore, it is important to confirm the acid resistance of the sample. As shown in Fig. S5, the ECH chemically cross-linked P/Fe-CB-0.5 exhibited excellent acid resistance. It remained undissolved after 600 min of shaking even at pH 2.0. In contrast, the Fe-chitosan beads without using ECH cross-linkage were almost completely dissolved after shaking for 4 h at pH 2. The concentration of dissolved Fe from samples at different pHs is illustrated in Fig. S6. Clearly, the ECH cross-linkage significantly inhibited the dissolution of ferrihydrite of the beads and almost no leaching of Fe was detected at pH above 1. The above observation demonstrated that ECH cross-linkage could significantly increase the acid resistance of the adsorbent and inhibit the dissolution of iron oxide. Thus, the prepared adsorbent can be safely used in water treatment process.
2.6. Kinetics and isotherms models The pseudo-first-order (Eq. 1) and the pseudo-second-order (Eq. 2) are presented for the adsorption kinetics.
qt = qe (1
exp( k1 t ))
(1)
qt = qe 1
1 1 + qe k2 t
(2)
where qe and qt (mg/g) are the amount of arsenic adsorbed at equilibrium and at time t (min), respectively; k1 (min−1) and k2 (g/mg/min) are the related adsorption rate constants. Langmuir (Eq. 3) and Freundlich (Eq. 4) models are presented for the adsorption isotherms:
qe =
qe =
qmax KL Ce 1 + KL Ce 1 KF Ce n
(3) (4)
where Ce (mg/L) is the equilibrium arsenic concentration in the solution, qmax (mg/g) is the maximum adsorption capacity, KL is the Langmuir adsorption constant, and KF and n are the Freundlich adsorption constants.
3.2. Adsorption kinetics
3. Results and discussion
Fig. 2 revealed the adsorption kinetics of As(III) on P/Fe-CB and FeCB at two different initial As(III) concentrations. As(III) adsorption onto P/Fe-CB was significantly faster than that of Fe-CB during the whole adsorption process. For P/Fe-CB, adsorption equilibrium reached in approximately 240 min for 0.975 mg/L As(III) (Fig. 2a). In contrast, it took more than 600 min for Fe-CB to reach adsorption equilibrium under the same experimental conditions. A similar trend was observed at an initial As(III) concentration of 2.71 mg/L (Fig. 2b). The fast adsorption of As(III) onto P/Fe-CB was ascribed to its unique pore structure, which favored the diffusion of As(III) from the solution to the inner adsorption sites of P/Fe-CB. The distribution of As(III) in the As
3.1. Characterizations The undried P/Fe-CB-0.5 precursors were uniform reddish-brown beads with a diameter of ˜ 2.0 mm (Fig. S2). After freeze-drying, the diameter of P/Fe-CB-0.5 beads was reduced to ˜ 1.0 mm (Fig. 1a). The SEM image showed that the surface of P/Fe-CB-0.5 was full of nanometer-scale pores (Fig. 1b). Significantly, the cross-sectional view of the bead showed that its interior possessed larger radially aligned micron-sized tunnels from outside to inside (Fig. 1c). Similarly, the P/Fe-
Fig. 1. Photograph (a), surface SEM image (b) and cross-section SEM image (c) of P/Fe-CB-0.5. Photograph (d), surface SEM image (e) and cross-section SEM image (f) of Fe-CB-0.5. 99
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Fig. 2. As(III) adsorption kinetics on P/Fe-CB and Fe-CB: with initial As(III) concentration of 0.975 mg/L (a) and 2.71 mg/L (b). (c) EDS line spectrum of As-L on the cross-section of P/Fe-CB-0.5. Adsorbent dose = 1 g/L, pH = 7, T = 25℃, C0As(III) = 10 mg/L (c), and t = 24 h (c).
(III)-loaded P/Fe-CB-0.5 was investigated by EDS (Fig. 2c). The adsorbed As(III) was distributed throughout the bead, demonstrating that As(III) was successfully impregnated into the adsorbent. Note that chitosan itself showed no adsorption of As(III) (data not presented), indicating that the adsorption of As(III) onto P/Fe-CB and Fe-CB was ascribed to the loaded ferrihydrite. The kinetics data of As(III) adsorption onto P/Fe-CB and Fe-CB were fitted to pseudo-first-order and pseudo-second-order models. The rate constants are summarized in Table S2. The adsorption could be better described by the pseudo-second order model than pseudo-first-order model for all the samples, which indicated that the chemical adsorption may be the potential rate-limiting step in the arsenic adsorption [33,34]. The adsorption rate constants (k1 and k2) of As(III) adsorption onto P/Fe-CB were much higher than that onto Fe-CB under the similar experimental conditions, indicative of a faster removal of As(III) by P/ Fe-CB than by Fe-CB. In addition, Fe content in both P/Fe-CB and Fe-CB had no significant effect on As(III) removal in the studied As(III) concentration range. Thus, only P/Fe-CB-0.5 and P/Fe-CB-3.0 were further investigated.
and within 120 min for P/Fe-CB-3.0 when 0.5 mM H2O2 was added. 3.4. Adsorption isotherms The adsorption isotherms of As(III) by P/Fe-CB-0.5 and P/Fe-CB-3.0 without H2O2 are presented in Fig.4a. For Fe-CB-3.0, the As(III) adsorption capacity is larger than 4.65 mg/g at equilibrium concentration below 10 μg/L. Both Langmuir and Freundlich models were used to describe the adsorption isotherms. The parameters are listed in Table 1. The higher correlation coefficient (R2) indicated that the Freundlich model could be more suitable than the Langmuir model for describing the adsorption process, which was possibly due to the nonuniform active sites and the irreversible adsorption [21]. The maximal adsorption capacities calculated from the Langmuir model for P/Fe-CB-0.5 and P/ Fe-CB-3.0 were 27.8 and 57.2 mg/L, respectively. The higher adsorption capacity of P/Fe-CB-3.0 than P/Fe-CB-0.5 should be attributed to its higher Fe content which provided more adsorption sites. The adsorption capacities of P/Fe-CB-3.0 for As(III) is outperforming most of iron/chitosan-based adsorbents (Table S3). It is also much higher than many commercial arsenic adsorbents such as granular ferric hydroxide (GFH) (8 mg/g) and Bayoxide® E-33 (7.56–12.86 mg/g) [35,36]. Fig. 4b showed the adsorption isotherms of As(III) on P/Fe-CB-3.0 and P/Fe-CB-0.5 in the presence of H2O2 (As(III) / H2O2 = 1 mg/L / 0.5 mM). The addition of H2O2 significantly reduced the corresponding equilibrium arsenic concentration, especially for P/Fe-CB-3.0. It further confirmed the promotion of arsenite adsorption by H2O2.
3.3. Effect of H2O2 on As(III) removal Fig. 3a and b showed the effect of H2O2 on As(III) removal by P/FeCB-0.5 and P/Fe-CB-3.0, respectively. Impressively, the removal rate of As(III) was accelerated when a small amount of H2O2 was added and the residual arsenic in solution was below 10 μg/L. Taking the strict MCL standard into consideration, it is very meaningful for its practical applications. A similar result was observed for the adsorption of As(III) with a higher initial As(III) concentration of 2.71 mg/L (Fig. S7). Such enhancement in As(III) removal possibly resulted from the oxidation of As(III). The removal rate increased as the H2O2 concentration increased from 0.1 to 2 mM. Considering the effectiveness and cost, 0.5 mM was chosen as the feasible H2O2 concentration because the arsenic concentration dropped below to 10 μg/L within 240 min for P/Fe-CB-0.5
3.5. Solution chemistry 3.5.1. Effect of pH The effect of pH on As(III) removal by P/Fe-CB-3.0 and P/Fe-CB-0.5 was elucidated in Fig. 5a. The corresponded removal kinetics are depicted in Fig. S8. In the absence of H2O2, the As(III) adsorption on both P/Fe-CB-3.0 and P/Fe-CB-0.5 was above 98% at pH < 9. When the pH
Fig. 3. Effect of H2O2 on As(III) removal by P/Fe-CB-0.5 (a) and P/Fe-CB-3.0 (b). C0As(III) = 0.975 mg/ L, adsorbent dose = 1 g/L, pH = 7, and T = 25℃. 100
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Fig. 4. Adsorption isotherms of As(III) by P/Fe-CB-0.5 and P/Fe-CB-3.0 without H2O2 (a) and with H2O2 (b). Adsorbent dose = 1 g/L, pH = 7, As(III) / H2O2 = 1 mg/L / 0.5 mM for (b), t = 72 h, and T = 25℃.
adsorption. Comparatively, the As(III) was completely removed by P/ Fe-CB-3.0 and P/Fe-CB-0.5 at pH < 9 when H2O2 was added. The addition of H2O2 accelerated the removal of As(III) at pH < 9 (Fig. S8). However, further increasing pH from 9 to 11, the As(III) removal efficiency declined sharply to 53.7% and 32.6% for P/Fe-CB-3.0 and P/FeCB-0.5, respectively, which was much lower than that in the absence of H2O2. The possible reason was that H2O2 oxidized As(III) to As(V) which existed in the more negatively charged species of HAsO42- and AsO43- at pH 11. Thereby, enhanced electrostatic repulsion reduced the arsenic removal. The results indicated that the addition of H2O2 was not suitable for the treatment of highly alkaline As(III)-contaminated wastewater (pH > 9) by the prepared adsorbent, and this aspect needs to be considered prior to application of H2O2 for As(III) removal in real natural waters.
Table 1 Langmuir and Freundlich isotherm parameters for As(III) adsorption. sample
P/Fe-CB-0.5 P/Fe-CB-3.0
Langmuir
Freundlich
qmax (mg/g)
KL (L/mg)
R2
KF (mg/g)
n
R2
27.8 52.7
0.348 0.694
0.9136 0.9604
9.86 20.88
3.23 2.35
0.9862 0.9849
was increased to 11, the removal efficiency decreased but still maintained above 89%. The point of zero charge (pHPZC) of P/Fe-CB-0.5 and P/Fe-CB-3.0 was about 8.7 and 7.7, respectively (Fig. S9). For pH values above 9.2, the As(III) deprotonated into negatively charged anion species of H2AsO3−. On the other hand, the adsorbent was negatively charged. Thus, physical absorption would be inhibited by the electrostatic repulsion between the negatively charged adsorbent and As(III) anions, and the adsorption of As(III) was mainly attributed to chemical
3.5.2. Effect of coexisting anions Fig. 5b and c show the commonly present anions on As(III) removal
Fig. 5. Effect of pH (a), anions (b, c), and HA (d) on the As(III) removal by P/Fe-CB-0.5 and P/Fe-CB-3.0 with/without H2O2. C0As(III) = 0.975 mg/L, adsorbent dose = 1 g/L, pH = 7 (for b, c and d), t = 12 h, and T = 25℃. 101
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by P/Fe-CB-0.5 and P/Fe-CB-3.0, respectively. The removal kinetics are depicted in Fig. S10. The presence of bicarbonate (CO32−) and sulfate (SO42−) did not influence the As(III) removal in the studied concentration ranges. Additionally, the addition of H2O2 also promoted the removal of As(III). Generally, silicate (SiO32−) and phosphate (PO43-) are regarded as the strongly interfering substances for As(III) and As(V) removal because they could also form the inner-sphere complex with metal oxides [37]. Only a slight reduction in As(III) removal was observed in the presence of silicate. Additionally, the silicate, even at concentrations up to 50 mg/L, had an insignificant effect on As(III) removal when the H2O2 was added. The influence of phosphate on As (III) removal was more serious than that of silicate, especially in the presence of H2O2. This was mainly due to the extraordinary chemical similarity of phosphate and arsenate. Note that typical concentrations of phosphate in groundwater are 0.02-0.6 mg/L [38,39]. The removal efficiency was only slightly decreased when the level of phosphates in the solution below 5 mg/L. Therefore, the P/Fe-CB is desirable for its application in groundwater.
with the FTIR results. The surface hydroxyl groups (eOH) are known to play a key role in arsenic adsorption. There are three possible inner-sphere arsenic complexes resulting from arsenic interaction with eOH: bidentate binuclear, bidentate mononuclear and monodentate mononuclear complexes (Fig. S13) [43]. The O1s XPS spectra are shown in Fig. 6d and the fitted data are listed in Table S4. The stoichiometric ratio of surface hydroxyl between the fresh adsorbent and the adsorbent saturated with arsenic was 1:2 for a monodentate surface complex and 2:1 for a bidentate one. The proportion of OH−/O2- increased from 1.06 to 1.31 and 1.46 after As(III) and As(III)/H2O2 adsorption, respectively. The increase in OH−/O2- ratio indicated that the arsenic adsorption was mainly through the formation of the monodentate mononuclear complex [44,45]. 3.6.4. As(III) oxidation mechanism The oxidation of As(III) by H2O2 itself at pH 7 was negligible (Fig. S14). Therefore, other reactive species existed in the P/Fe-CB/H2O2 system for As(III) oxidation. Previous studies reported that H2O2 could react with the surface ^Fe(III) of iron oxides to produce superoxide radicals (O2%ˉ) and surface ^Fe(II), and the ^Fe(II) could further react with H2O2 to produce hydroxyl radicals (%OH) [46,47]. ESR was performed to detect radicals (Fig. S15). Both DMPO-%OH and DMPO-O2%ˉ signals were not found in the P/Fe-CB/H2O2 system, indicating that these radicals were not formed. Furthermore, the addition of tert-butanol (TBA) as %OH scavenger and superoxide dismutase (SOD) as O2%ˉ scavenger had no effect on As(III) removal (Fig. S16). Consequently, the % OH and O2%ˉ were excluded for As(III) oxidation. Some studies reveal that high-valent iron species (^FeIV]O) are the dominant oxidant at near neutral pH. High-valent iron species could specifically oxidize dimethyl sulfoxide (DMSO) to dimethyl sulfone (DMSO2) [48,49]. The consumption of DMSO during the reaction process was monitored. However, its consumption was negligible regardless of the presence of H2O2 (Fig. S17). Additionally, DMSO2 was not detected by GC–MS in the P/Fe-CB/H2O2 system (Fig. S18). Furthermore, the addition of DMSO did not inhibit As(III) removal (Fig. S19). These results demonstrated that the high-valent iron species were not formed in the P/ Fe-CB/H2O2 system. Numerous studies have proposed that the H2O2induced surface peroxide-like species formed on the oxides act as the dominant active species for As(III) oxidation [30,50]. The diffuse UV–vis reflectance spectra of H2O2-treated P/Fe-CB-3.0 occurred red shift compared with the virgin one (Fig. S20), indicative of the formation of peroxide-like species on the H2O2-treated metal oxide [30]. Kim et al. [30] reported that p-benzoquinone (BQ) could be used as peroxide-like species scavenger. As shown in Fig. S21, the addition of BQ significantly inhibited the removal of As(III), indicating that the H2O2-induced surface peroxide-like species were the main oxidant species for As(III) oxidation.
3.5.3. Effect of humic acid Humic acid (HA) widely exists in natural water. The effect of HA on the As(III) removal is showed in Fig. 5d. The corresponded removal kinetics are depicted in Fig. S11. The presence of HA did not have any influence on As(III) removal even at HA concentration up to 20 mg/L. Moreover, the addition of H2O2 promoted As(III) removal by P/Fe-CB0.5 and P/Fe-CB-3.0. The results suggested that P/Fe-CB could effectively remove arsenic in the absence or presence of HA. 3.6. As(III) adsorption mechanism 3.6.1. Effect of ionic strength The increase of ionic strength is known to significantly influence outer-sphere interactions between solute and adsorbent surface but not affect inner-sphere complexation [40]. The solution ionic strength was adjusted by the addition of NaCl. As shown in Fig. S12, the ionic strength showed a negligible effect on As(III) removal in the absence of H2O2, indicating that the As(III) adsorption onto the P/Fe-CB followed the inner-sphere complex mechanism. In contrast, the presence of NaCl only slightly suppressed the As(III) removal kinetics when H2O2 was added, and the inhibition was negligible at the end of the reaction. Thereby, the addition of H2O2 did not significantly change the arsenic adsorption mechanism. 3.6.2. FTIR analysis To further elucidate the As(III) adsorption mechanism, the FTIR characterizations of P/Fe-CB-3.0 before and after As(III) adsorption were performed (Fig. 6a). A new peak at 785 cm−1 appeared after As (III) adsorption. At pH 7, the dominant arsenite species in aqueous solution was As(OH)3 which had no IR bands in the 750-800 cm−1 region [41]. Therefore, the 785 cm−1should result from the formation of an inner-sphere surface complex. In the presence of H2O2, a new peak at 835 cm−1 appeared, which was ascribed to the As-O stretching in AsO-Fe linkage between As(V) and ferrihydrite [42]. This result indicated that: (1) the As(III) adsorbed on P/Fe-CB-3.0 was eventually oxidized to As(V) when H2O2 was added, and (2) both As(III) and As(V) adsorption on P/Fe-CB-3.0 followed the inner-sphere complex mechanism.
3.7. Performances of P/Fe-CB in different water matrices Natural water matrices are considerably more complex than deionized water because they contain various inorganic ions and organic matters. To better assess the feasibility of P/Fe-CB to As(III) removal under environmentally relevant conditions, three types of As(III)-spiked natural water matrices (groundwater, river water and lake water) were used. The water quality parameters are shown in Table S5. Fig. 7 showed the As(III) removal by P/Fe-CB-3.0 in the three water matrices. In all cases, the equilibrium As(III) concentration dropped from 200 μg/ L to below 10 μg/L within 360 min (Fig. 7a). When 0.1 mM H2O2 was added, the process could be shortened to 240 min (Fig. 7b). Similar results were obtained for P/Fe-CB-0.5 (Fig. S22). Thus, 1 L of As(III)contaminated natural water (200 μg/L) could be treated to meet the WHO standard for arsenic in drinking water using only 1 g of P/Fe-CB.
3.6.3. XPS analysis XPS measurements were performed to determine the surface chemistry of the adsorbents. As shown in Fig. 6b, new As3d as well as AsLMM peaks appeared after reaction with As(III) or As(III)/H2O2, suggesting that the arsenic was successfully bonded to P/Fe-CB-3.0. To reveal the arsenic species on P/Fe-CB-3.0, the As3d XPS spectra were analyzed (Fig. 6c). In the absence of H2O2, almost all the arsenic adsorbed on P/Fe-CB-3.0 was in the form of As(III). Comparatively, it was completely transformed to As(V) when H2O2 was added, agreeing well 102
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Fig. 6. FT-IR (a) and XPS (b) spectra of the P/Fe-CB-3.0 before and after As(III) or As(III)/H2O2 adsorption, (c) As3d XPS spectra, and (d) O1s XPS spectra.
3.8. Adsorbent regeneration and reuse
synthetic As(III)-contaminated groundwater containing As(III) and competing anions (CO32−, SO42−, SiO32−, PO43-). Their breakthrough curve is depicted in Fig. 8b. The P/Fe-CB-3.0 column generated about 3000 BV (27.12 L) effluent before the breakthrough point (10 μg/L) occurred. Additionally, the total Fe concentration leaching from the column was below 0.05 mg/L. The column arsenite adsorption density (qcolumn) at the breakthrough point was 1.19 mg/g, which was lower than the adsorption capacity from batch tests. This was due to the fact that the low initial As(III) concentration and very short contact time (4.5 min) of column operation made less As(III) be captured by P/FeCB-3.0. Thus, P/Fe-CB-3.0 was effective for As(III) removal from solution using bath batch or column operation.
Recyclability a crucial aspect to evaluate adsorbent applicability. Fig. 8a showed the successive adsorption − desorption cyclic runs of As (III) by P/Fe-CB-0.5 and P/Fe-CB-3.0. After each adsorption cycle, the adsorbent was regenerated with 0.1 M NaOH. In the absence of H2O2, the removal efficiency of As(III) reduced slightly as the regeneration cycle increased. After recycling four times, the removal efficiency of As (III) still maintained above 90%. More satisfactory results were achieved when 0.5 mM H2O2 was added and the removal efficiency was maintained at nearly 100% throughout five consecutive cycles. 3.9. Column operation of P/Fe-CB Fixed-bed adsorption was carried out to remove arsenite from
Fig. 7. As(III) removal in natural water samples by P/Fe-CB-3.0 (a) and in the presence of H2O2 (b). C0As(III) = 0.2 mg/L, adsorbent dose = 1 g/L, pH = 7, H2O2 = 0.1 mM (if added), and T = 25℃. 103
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Fig. 8. (a) As(III) removal in 5 consecutive cycles for P/Fe-CB, and (b) As(III) removal from a synthetic groundwater by using separate columns packed with P/Fe-CB3.0. C0As(III) = 0.975 mg/L for (a), adsorbent dose = 1 g/L, pH = 7, t = 12 h, and T = 25℃.
4. Conclusions
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Novel porous Fe-chitosan beads (P/Fe-CB) were prepared by the freeze-casting technique. The P/Fe-CB possessed unique pore structure and excellent acid resistance. The maximal adsorption capacities of P/ Fe-CB for As(III) was 52.7 mg/g at pH 7.0, which was higher than that for most of the reported granular adsorbents. Coexisting sulfate, carbonate, silicate and humic acid had no significant effect on As(III) removal, especially when H2O2 was added. The P/Fe-CB could be readily separated and regenerated, and the adsorption efficiency maintained above 90% throughout five consecutive cycles. For most situation, the addition of H2O2 accelerated the As(III) removal. In the presence of high concentration of phosphate (50 mg/L) or under high pH (11) condition, the addition of H2O2 would inhibit the As(III) removal. Moreover, it could effectively remove As(III) from natural water matrices. The fixed-bed working capacity was about 3000 BV. Therefore, the P/Fe-CB could be employed as a promising adsorbent for the As(III) removal from water. Acknowledgments This work was supported by the National Natural Science Foundation of China (51778218 and 51478171) and Technology Innovation Plan of Hunan Province (2017SK2420). Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.jhazmat.2019.03.070. References [1] D.K. Nordstrom, Worldwide occurrences of arsenic in ground water, Science 296 (2002) 2143–2145, https://doi.org/10.1126/science.1072375. [2] L. Winkel, M. Berg, M. Amini, S.J. Hug, C.A. Johnson, Predicting groundwater arsenic contamination in Southeast Asia from surface parameters, Nat. Geosci. 1 (2008) 536–542, https://doi.org/10.1038/ngeo254. [3] M. Argos, T. Kalra, P.J. Rathouz, Y. Chen, B. Pierce, F. Parvez, T. Islam, A. Ahmed, M. Rakibuz-Zaman, R. Hasan, Arsenic exposure from drinking water, and all-cause and chronic-disease mortalities in Bangladesh (HEALS): a prospective cohort study, Lancet 376 (2010) 252–258, https://doi.org/10.1016/S0140-6736(10)60481-3. [4] P. Smedley, D. Kinniburgh, A review of the source, behaviour and distribution of arsenic in natural waters, Appl. Geochem. 17 (2002) 517–568, https://doi.org/10. 1016/S0883-2927(02)00018-5. [5] C. Su, Environmental implications and applications of engineered nanoscale magnetite and its hybrid nanocomposites: a review of recent literature, J. Hazard. Mater. 322 (2017) 48–84, https://doi.org/10.1016/j.jhazmat.2016.06.060. [6] S. Dixit, J.G. Hering, Comparison of arsenic (V) and arsenic (III) sorption onto iron oxide minerals: implications for arsenic mobility, Environ. Sci. Technol. 37 (2003) 4182–4189, https://doi.org/10.1021/es030309t. [7] X. Zhang, C. Cheng, J. Qian, Z. Lu, S. Pan, B. Pan, Highly efficient water decontamination by using sub-10 nm FeOOH confined within millimeter-sized mesoporous polystyrene beads, Environ. Sci. Technol. 51 (2017) 9210–9218, https://doi.org/
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Enhanced arsenite removal from water by radially porous Fe-chitosan beads: Adsorption and H2O2 catalytic oxidation
T
Yuanfeng Weia, Xingwen Yua, Chengbin Liua, , Jianhong Mab, Shudan Weia, Tao Chena, Kai Yina, Hui Liua, Shenglian Luoa ⁎
a b
State Key Laboratory of Chemo/Biosensing and Chemometrics, Hunan University, Changsha, 410082, PR China College of Environmental Science and Engineering, Hunan University, Changsha, 410082, PR China
GRAPHICAL ABSTRACT
ARTICLE INFO
ABSTRACT
Keywords: Fe-chitosan bead H2O2 Arsenite Adsorption Fix-bed process
Although Fe-chitosan adsorbents are attractive for removing arsenite from water, the practical applications of these granular adsorbents are mainly limited by slow adsorption kinetics. In this study, radially porous Fechitosan beads (P/Fe-CB) were prepared using freeze-casting technique. The P/Fe-CB particles possess radially aligned micron-sized tunnels from the surface to the inside as well as excellent acid resistance. Kinetic studies show that the adsorption equilibrium time of P/Fe-CB to 0.975 mg/L As(III) (within 240 min) is considerably shorter than that of compact Fe-chitosan beads (over 600 min). The maximal adsorption capacity of P/Fe-CB for As(III) is 52.7 mg/g. It can work effectively in a wide pH range from 3 to 9, and the coexisting sulfate, carbonate, silicate and humic acid have no significant effect on As(III) removal. The addition of H2O2 can further accelerate and promote the As(III) removal except at high pH (11) and phosphate concentration (50 mg/L). The fixed-bed experiments demonstrate that the P/Fe-CB column can effectively treat about 3000 bed volume (BV) of simulated As(III)-containing groundwater to meet the drinking water standard (< 10 μg As/L). This study would extend the potential applicability of porous Fe based chitosan adsorbent and millimeter-sized adsorbent combined with H2O2 to a great extent.
⁎
Corresponding author. E-mail address: [email protected] (C. Liu).
https://doi.org/10.1016/j.jhazmat.2019.03.070 Received 18 December 2018; Received in revised form 26 February 2019; Accepted 16 March 2019 Available online 18 March 2019 0304-3894/ © 2019 Elsevier B.V. All rights reserved.
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1. Introduction
with H2O2 for As(III) removal.
Arsenic as a carcinogenic element has become a global threat to the surface water and groundwater due to the natural and anthropogenic release [1]. Arsenic contaminated water is served as the primary drinking water source in many regions [2,3]. To minimize the risk of arsenic, the World Health Organization (WHO) lowered the maximum contaminant level (MCL) of arsenic in drinking water to 10 μg L−1. Arsenic exists mostly in two inorganic forms as oxyanions of arsenite As (III) and arsenate As(V) in natural groundwater. The dominant arsenic species in groundwater is As(III), which is 25˜60 times more toxic than As(V) [4]. Hence, removal of As(III) from the affected water is crucial. Iron (oxyhydr)oxides have been extensively used for arsenic removal, owing to their promising properties such as high sorption affinity toward arsenic, low cost, rich source, and environmental benignity [5,6]. Iron (oxyhydr)oxides usually appear as fine or ultrafine powders which are difficult to apply to fixed-bed columns due to the excessive pressure drop. To achieve a wider range of applications, iron (oxyhydr) oxides are usually immobilized on a support to eliminate this defect [7–9]. Fe-chitosan beads for arsenic removal hold considerable promise due to its practical advantages [10–12]. Chitosan is an abundant waste product that is derived from insect and shellfish exoskeletons [13]. It possesses many advantages such as inexpensive, biodegradable, and non-toxic [14]. Moreover, it can be easily formulated into beads. Despite extension in iron (oxyhydr)oxides application, Fe-chitosan beads usually shown slow kinetics due to its low porosity. While, practical column filtration typically has a short empty bed contact time (EBCT) [15]. Creation of porous chitosan beads is a workable solution to speed up the internal diffusion, thereby speeding up reaction kinetics. Moreover, this operation makes more adsorption sites accessible and thus improvement in adsorption capacity. Freeze-casting is an effortless technique of making porous structure, which exploits endogenous ice crystals as templates to shape and press building blocks to achieve a porous structure [16–18]. This technique does not require additional reagents. Moreover, the resulted pores are aligned [19], which can reduce the ionic diffusion path, facilitate ionic motion to the inner part of the bead, and improve utilization of adsorption sites. In addition to improving the adsorbent, oxidation of uncharged As (III) to As(V) anion is an important strategy to accelerate the As(III) removal as well as increase the As(III) adsorption capacity, because As (V) has a stronger affinity to most of adsorbents relative to As(III) [6,20–24]. Among various oxidants, hydrogen peroxide (H2O2) is the most benign oxidant because its end product is H2O. However, the oxidation activity of H2O2 is strongly pH-dependent and limited to alkaline condition [25]. Integration of H2O2 with iron oxides can generate powerful reactive species even under acidic conditions [26,27]. Despite all this, only few studies focus on metal oxide/H2O2 for As(III) removal [28–30]. Additionally, these studies are limited to nanosized metal oxide and batch experiments. Furthermore, the comparison of As (III) removal by metal oxide with or without H2O2 has not been systematically investigated in the presence of coexisting ions and at different pH conditions. Herein, we fabricated a porous Fe-chitosan bead (P/Fe-CB) using a simple freeze-casting technique. Additionally, the P/Fe-CB was chemically cross-linked by epichlorhydrine (ECH) to markedly improve its acid resistance, which is usually ignored for Fe-based adsorbents. This P/Fe-CB possesses some advantages: fast adsorption, high adsorption capacity and excellent acid resistance. A heterogeneous catalytic oxidation process was also developed by integrating P/Fe-CB and H2O2 to further promote and accelerate As(III) removal. The effect of H2O2 on As(III) removal by P/Fe-CB was systematically investigated. Both batch and column experiments were carried out to examine its adsorption characteristics and to evaluate the potential for practical application. This work would open up a new avenue for the construction of porous Fe based adsorbents and the application of millimeter-sized adsorbents
2. Materials and methods 2.1. Chemicals All chemicals were analytical grade. Chitosan powder (deacetylation: 95%) was purchased from Aladdin Chemistry Co., Ltd. Ferric chloride (FeCl3⋅6H2O) and epichlorohydrin (ECH) were purchased from Sinopharm Chemical Reagent Co., Ltd., Shanghai, China. As(III) stock solution was prepared with NaOH solution using As2O3. As(III) working solution was freshly prepared by diluting arsenic stock solutions with deionized water. 2.2. Preparation of porous Fe-chitosan beads 1 g of chitosan was completely dissolved in 30 mL of acetic acid solution (2% (v/v)). Then, 4 mL of 0.46, 0.92, 1.85 or 2.77 M FeCl3 solution was added into the solution and stirred until it became homogenous. 1 mL of ECH was then added into the homogenous solution under vigorous agitation. The resulting solution was pushed through a 10-mL syringe fitted with 8 G needles into 0.5 M NaOH solution to form beads. The resultant beads were kept for 24 h in the NaOH bath and then washed with deionized water until the residual water was neutral. The wet beads were placed in an ultra-low temperature freezer (−71 °C) for 1 d. After that, these beads dried by freeze-drying. The beads are abbreviated as P/Fe-CB-X (X = 0.5, 1, 2, 3), where X stands for the mass ratio of FeCl3 / chitosan. For comparison, the samples naturally dried at room temperature was prepared and abbreviated as Fe-CB-X (X = 0.5, 1, 2, 3). 2.3. Characterizations The morphologies of the samples were observed on a Hitachi S-4800 scanning electron microscope operating at 3.0 kV. The crystal phases of the samples were collected on an X-ray diffractometer with Cu-Kα radiation (XRD, M21X, MAC Science Ltd.). The surface chemistry properties of the adsorbent before and after the adsorption process were determined by X-ray photoelectron spectroscopy (XPS, K-α 1063, Thermo Fisher Scientific) and Fourier transform infrared spectroscopy (FTIR, Bruker TENSOR 27). The pH values at the point of zero charge (pHPZC) of the adsorbent were measured in a series of 0.01 M NaCl with different initial pH. The Brunner − Emmet − Teller (BET) specific surface areas were measured on a Belsorp-Mini II analyser. 2.4. Batch adsorption experiments Most experiments were carried out in a 50-mL glass bottle containing 30 mL of 0.975 mg/L arsenite solution. During the adsorption process, the reaction system was shaken in a thermostatic orbit shaker at 180 rpm under 25 °C. The adsorbent dose was 1 g/L and the initial pH values of the arsenic solutions were adjusted with HCl solution or NaOH solution. The ionic strength in the solution was adjusted by the addition of NaCl. The effect of humic acid and ions was examined by adding preset amount of humic acid, Na2CO3, Na2SO4, Na2SiO3 and Na3PO4 into the test solution. To make the pH constant during the reaction, moderate amount of NaOH or HCl was added every three hours. The concentration of residual arsenic was determined by an atomic fluorescence spectrophotometer (AFS-9700, Beijing Haiguang Instrument Inc., China). 2.5. Column experiments Column adsorption experiments were carried out at ambient temperature using P/Fe-CB (4.3 g) packed in fixed-bed columns (12 mm in diameter and 100 mm in length). The schematic flowchart is presented 98
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in Fig.S1 (Supporting Information). About 10 mm thick of silica wool was used at each end of the column to distribute the flow. The effective bed volume (BV) of the column was 9.04 mL A peristaltic pump was used to maintain an upward flow. The empty bed contact time (EBCT) was set as 4.5 min and the corresponded flow rate was 2.0 mL/min.
CB-1.0, P/Fe-CB-2.0 and P/Fe-CB-3.0 samples possessed similar porous structures (Fig. S3). The characteristic porous structures would substantially accelerate the transfer rate of arsenite inside the adsorbent. In contrast, the naturally dried Fe-CB-0.5 exhibited a quite different surface and internal structure (Fig. 1d-f). The surface of Fe-CB-0.5 beads was rough and its inside was compact. The BET surface areas of all the sample are summarized in Table S1. The BET surface area of P/Fe-CB beads was much larger than that of Fe-CB beads. The XRD pattern of P/ Fe-CB-0.5 is shown in Fig. S4. The broad diffraction peak at 20.2° is ascribed to chitosan [31], while the peaks at approximately 35.1 and 62.2° correspond to the diffraction of the (110) and (115) planes of ordered 2-line ferrihydrite [32]. Chitosan and ferrihydrite (iron oxide) are readily soluble in dilute acidic solutions, while some of the As-contaminated wastewater are acidic, such as the acid mine drainage (AMD). Therefore, it is important to confirm the acid resistance of the sample. As shown in Fig. S5, the ECH chemically cross-linked P/Fe-CB-0.5 exhibited excellent acid resistance. It remained undissolved after 600 min of shaking even at pH 2.0. In contrast, the Fe-chitosan beads without using ECH cross-linkage were almost completely dissolved after shaking for 4 h at pH 2. The concentration of dissolved Fe from samples at different pHs is illustrated in Fig. S6. Clearly, the ECH cross-linkage significantly inhibited the dissolution of ferrihydrite of the beads and almost no leaching of Fe was detected at pH above 1. The above observation demonstrated that ECH cross-linkage could significantly increase the acid resistance of the adsorbent and inhibit the dissolution of iron oxide. Thus, the prepared adsorbent can be safely used in water treatment process.
2.6. Kinetics and isotherms models The pseudo-first-order (Eq. 1) and the pseudo-second-order (Eq. 2) are presented for the adsorption kinetics.
qt = qe (1
exp( k1 t ))
(1)
qt = qe 1
1 1 + qe k2 t
(2)
where qe and qt (mg/g) are the amount of arsenic adsorbed at equilibrium and at time t (min), respectively; k1 (min−1) and k2 (g/mg/min) are the related adsorption rate constants. Langmuir (Eq. 3) and Freundlich (Eq. 4) models are presented for the adsorption isotherms:
qe =
qe =
qmax KL Ce 1 + KL Ce 1 KF Ce n
(3) (4)
where Ce (mg/L) is the equilibrium arsenic concentration in the solution, qmax (mg/g) is the maximum adsorption capacity, KL is the Langmuir adsorption constant, and KF and n are the Freundlich adsorption constants.
3.2. Adsorption kinetics
3. Results and discussion
Fig. 2 revealed the adsorption kinetics of As(III) on P/Fe-CB and FeCB at two different initial As(III) concentrations. As(III) adsorption onto P/Fe-CB was significantly faster than that of Fe-CB during the whole adsorption process. For P/Fe-CB, adsorption equilibrium reached in approximately 240 min for 0.975 mg/L As(III) (Fig. 2a). In contrast, it took more than 600 min for Fe-CB to reach adsorption equilibrium under the same experimental conditions. A similar trend was observed at an initial As(III) concentration of 2.71 mg/L (Fig. 2b). The fast adsorption of As(III) onto P/Fe-CB was ascribed to its unique pore structure, which favored the diffusion of As(III) from the solution to the inner adsorption sites of P/Fe-CB. The distribution of As(III) in the As
3.1. Characterizations The undried P/Fe-CB-0.5 precursors were uniform reddish-brown beads with a diameter of ˜ 2.0 mm (Fig. S2). After freeze-drying, the diameter of P/Fe-CB-0.5 beads was reduced to ˜ 1.0 mm (Fig. 1a). The SEM image showed that the surface of P/Fe-CB-0.5 was full of nanometer-scale pores (Fig. 1b). Significantly, the cross-sectional view of the bead showed that its interior possessed larger radially aligned micron-sized tunnels from outside to inside (Fig. 1c). Similarly, the P/Fe-
Fig. 1. Photograph (a), surface SEM image (b) and cross-section SEM image (c) of P/Fe-CB-0.5. Photograph (d), surface SEM image (e) and cross-section SEM image (f) of Fe-CB-0.5. 99
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Fig. 2. As(III) adsorption kinetics on P/Fe-CB and Fe-CB: with initial As(III) concentration of 0.975 mg/L (a) and 2.71 mg/L (b). (c) EDS line spectrum of As-L on the cross-section of P/Fe-CB-0.5. Adsorbent dose = 1 g/L, pH = 7, T = 25℃, C0As(III) = 10 mg/L (c), and t = 24 h (c).
(III)-loaded P/Fe-CB-0.5 was investigated by EDS (Fig. 2c). The adsorbed As(III) was distributed throughout the bead, demonstrating that As(III) was successfully impregnated into the adsorbent. Note that chitosan itself showed no adsorption of As(III) (data not presented), indicating that the adsorption of As(III) onto P/Fe-CB and Fe-CB was ascribed to the loaded ferrihydrite. The kinetics data of As(III) adsorption onto P/Fe-CB and Fe-CB were fitted to pseudo-first-order and pseudo-second-order models. The rate constants are summarized in Table S2. The adsorption could be better described by the pseudo-second order model than pseudo-first-order model for all the samples, which indicated that the chemical adsorption may be the potential rate-limiting step in the arsenic adsorption [33,34]. The adsorption rate constants (k1 and k2) of As(III) adsorption onto P/Fe-CB were much higher than that onto Fe-CB under the similar experimental conditions, indicative of a faster removal of As(III) by P/ Fe-CB than by Fe-CB. In addition, Fe content in both P/Fe-CB and Fe-CB had no significant effect on As(III) removal in the studied As(III) concentration range. Thus, only P/Fe-CB-0.5 and P/Fe-CB-3.0 were further investigated.
and within 120 min for P/Fe-CB-3.0 when 0.5 mM H2O2 was added. 3.4. Adsorption isotherms The adsorption isotherms of As(III) by P/Fe-CB-0.5 and P/Fe-CB-3.0 without H2O2 are presented in Fig.4a. For Fe-CB-3.0, the As(III) adsorption capacity is larger than 4.65 mg/g at equilibrium concentration below 10 μg/L. Both Langmuir and Freundlich models were used to describe the adsorption isotherms. The parameters are listed in Table 1. The higher correlation coefficient (R2) indicated that the Freundlich model could be more suitable than the Langmuir model for describing the adsorption process, which was possibly due to the nonuniform active sites and the irreversible adsorption [21]. The maximal adsorption capacities calculated from the Langmuir model for P/Fe-CB-0.5 and P/ Fe-CB-3.0 were 27.8 and 57.2 mg/L, respectively. The higher adsorption capacity of P/Fe-CB-3.0 than P/Fe-CB-0.5 should be attributed to its higher Fe content which provided more adsorption sites. The adsorption capacities of P/Fe-CB-3.0 for As(III) is outperforming most of iron/chitosan-based adsorbents (Table S3). It is also much higher than many commercial arsenic adsorbents such as granular ferric hydroxide (GFH) (8 mg/g) and Bayoxide® E-33 (7.56–12.86 mg/g) [35,36]. Fig. 4b showed the adsorption isotherms of As(III) on P/Fe-CB-3.0 and P/Fe-CB-0.5 in the presence of H2O2 (As(III) / H2O2 = 1 mg/L / 0.5 mM). The addition of H2O2 significantly reduced the corresponding equilibrium arsenic concentration, especially for P/Fe-CB-3.0. It further confirmed the promotion of arsenite adsorption by H2O2.
3.3. Effect of H2O2 on As(III) removal Fig. 3a and b showed the effect of H2O2 on As(III) removal by P/FeCB-0.5 and P/Fe-CB-3.0, respectively. Impressively, the removal rate of As(III) was accelerated when a small amount of H2O2 was added and the residual arsenic in solution was below 10 μg/L. Taking the strict MCL standard into consideration, it is very meaningful for its practical applications. A similar result was observed for the adsorption of As(III) with a higher initial As(III) concentration of 2.71 mg/L (Fig. S7). Such enhancement in As(III) removal possibly resulted from the oxidation of As(III). The removal rate increased as the H2O2 concentration increased from 0.1 to 2 mM. Considering the effectiveness and cost, 0.5 mM was chosen as the feasible H2O2 concentration because the arsenic concentration dropped below to 10 μg/L within 240 min for P/Fe-CB-0.5
3.5. Solution chemistry 3.5.1. Effect of pH The effect of pH on As(III) removal by P/Fe-CB-3.0 and P/Fe-CB-0.5 was elucidated in Fig. 5a. The corresponded removal kinetics are depicted in Fig. S8. In the absence of H2O2, the As(III) adsorption on both P/Fe-CB-3.0 and P/Fe-CB-0.5 was above 98% at pH < 9. When the pH
Fig. 3. Effect of H2O2 on As(III) removal by P/Fe-CB-0.5 (a) and P/Fe-CB-3.0 (b). C0As(III) = 0.975 mg/ L, adsorbent dose = 1 g/L, pH = 7, and T = 25℃. 100
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Fig. 4. Adsorption isotherms of As(III) by P/Fe-CB-0.5 and P/Fe-CB-3.0 without H2O2 (a) and with H2O2 (b). Adsorbent dose = 1 g/L, pH = 7, As(III) / H2O2 = 1 mg/L / 0.5 mM for (b), t = 72 h, and T = 25℃.
adsorption. Comparatively, the As(III) was completely removed by P/ Fe-CB-3.0 and P/Fe-CB-0.5 at pH < 9 when H2O2 was added. The addition of H2O2 accelerated the removal of As(III) at pH < 9 (Fig. S8). However, further increasing pH from 9 to 11, the As(III) removal efficiency declined sharply to 53.7% and 32.6% for P/Fe-CB-3.0 and P/FeCB-0.5, respectively, which was much lower than that in the absence of H2O2. The possible reason was that H2O2 oxidized As(III) to As(V) which existed in the more negatively charged species of HAsO42- and AsO43- at pH 11. Thereby, enhanced electrostatic repulsion reduced the arsenic removal. The results indicated that the addition of H2O2 was not suitable for the treatment of highly alkaline As(III)-contaminated wastewater (pH > 9) by the prepared adsorbent, and this aspect needs to be considered prior to application of H2O2 for As(III) removal in real natural waters.
Table 1 Langmuir and Freundlich isotherm parameters for As(III) adsorption. sample
P/Fe-CB-0.5 P/Fe-CB-3.0
Langmuir
Freundlich
qmax (mg/g)
KL (L/mg)
R2
KF (mg/g)
n
R2
27.8 52.7
0.348 0.694
0.9136 0.9604
9.86 20.88
3.23 2.35
0.9862 0.9849
was increased to 11, the removal efficiency decreased but still maintained above 89%. The point of zero charge (pHPZC) of P/Fe-CB-0.5 and P/Fe-CB-3.0 was about 8.7 and 7.7, respectively (Fig. S9). For pH values above 9.2, the As(III) deprotonated into negatively charged anion species of H2AsO3−. On the other hand, the adsorbent was negatively charged. Thus, physical absorption would be inhibited by the electrostatic repulsion between the negatively charged adsorbent and As(III) anions, and the adsorption of As(III) was mainly attributed to chemical
3.5.2. Effect of coexisting anions Fig. 5b and c show the commonly present anions on As(III) removal
Fig. 5. Effect of pH (a), anions (b, c), and HA (d) on the As(III) removal by P/Fe-CB-0.5 and P/Fe-CB-3.0 with/without H2O2. C0As(III) = 0.975 mg/L, adsorbent dose = 1 g/L, pH = 7 (for b, c and d), t = 12 h, and T = 25℃. 101
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by P/Fe-CB-0.5 and P/Fe-CB-3.0, respectively. The removal kinetics are depicted in Fig. S10. The presence of bicarbonate (CO32−) and sulfate (SO42−) did not influence the As(III) removal in the studied concentration ranges. Additionally, the addition of H2O2 also promoted the removal of As(III). Generally, silicate (SiO32−) and phosphate (PO43-) are regarded as the strongly interfering substances for As(III) and As(V) removal because they could also form the inner-sphere complex with metal oxides [37]. Only a slight reduction in As(III) removal was observed in the presence of silicate. Additionally, the silicate, even at concentrations up to 50 mg/L, had an insignificant effect on As(III) removal when the H2O2 was added. The influence of phosphate on As (III) removal was more serious than that of silicate, especially in the presence of H2O2. This was mainly due to the extraordinary chemical similarity of phosphate and arsenate. Note that typical concentrations of phosphate in groundwater are 0.02-0.6 mg/L [38,39]. The removal efficiency was only slightly decreased when the level of phosphates in the solution below 5 mg/L. Therefore, the P/Fe-CB is desirable for its application in groundwater.
with the FTIR results. The surface hydroxyl groups (eOH) are known to play a key role in arsenic adsorption. There are three possible inner-sphere arsenic complexes resulting from arsenic interaction with eOH: bidentate binuclear, bidentate mononuclear and monodentate mononuclear complexes (Fig. S13) [43]. The O1s XPS spectra are shown in Fig. 6d and the fitted data are listed in Table S4. The stoichiometric ratio of surface hydroxyl between the fresh adsorbent and the adsorbent saturated with arsenic was 1:2 for a monodentate surface complex and 2:1 for a bidentate one. The proportion of OH−/O2- increased from 1.06 to 1.31 and 1.46 after As(III) and As(III)/H2O2 adsorption, respectively. The increase in OH−/O2- ratio indicated that the arsenic adsorption was mainly through the formation of the monodentate mononuclear complex [44,45]. 3.6.4. As(III) oxidation mechanism The oxidation of As(III) by H2O2 itself at pH 7 was negligible (Fig. S14). Therefore, other reactive species existed in the P/Fe-CB/H2O2 system for As(III) oxidation. Previous studies reported that H2O2 could react with the surface ^Fe(III) of iron oxides to produce superoxide radicals (O2%ˉ) and surface ^Fe(II), and the ^Fe(II) could further react with H2O2 to produce hydroxyl radicals (%OH) [46,47]. ESR was performed to detect radicals (Fig. S15). Both DMPO-%OH and DMPO-O2%ˉ signals were not found in the P/Fe-CB/H2O2 system, indicating that these radicals were not formed. Furthermore, the addition of tert-butanol (TBA) as %OH scavenger and superoxide dismutase (SOD) as O2%ˉ scavenger had no effect on As(III) removal (Fig. S16). Consequently, the % OH and O2%ˉ were excluded for As(III) oxidation. Some studies reveal that high-valent iron species (^FeIV]O) are the dominant oxidant at near neutral pH. High-valent iron species could specifically oxidize dimethyl sulfoxide (DMSO) to dimethyl sulfone (DMSO2) [48,49]. The consumption of DMSO during the reaction process was monitored. However, its consumption was negligible regardless of the presence of H2O2 (Fig. S17). Additionally, DMSO2 was not detected by GC–MS in the P/Fe-CB/H2O2 system (Fig. S18). Furthermore, the addition of DMSO did not inhibit As(III) removal (Fig. S19). These results demonstrated that the high-valent iron species were not formed in the P/ Fe-CB/H2O2 system. Numerous studies have proposed that the H2O2induced surface peroxide-like species formed on the oxides act as the dominant active species for As(III) oxidation [30,50]. The diffuse UV–vis reflectance spectra of H2O2-treated P/Fe-CB-3.0 occurred red shift compared with the virgin one (Fig. S20), indicative of the formation of peroxide-like species on the H2O2-treated metal oxide [30]. Kim et al. [30] reported that p-benzoquinone (BQ) could be used as peroxide-like species scavenger. As shown in Fig. S21, the addition of BQ significantly inhibited the removal of As(III), indicating that the H2O2-induced surface peroxide-like species were the main oxidant species for As(III) oxidation.
3.5.3. Effect of humic acid Humic acid (HA) widely exists in natural water. The effect of HA on the As(III) removal is showed in Fig. 5d. The corresponded removal kinetics are depicted in Fig. S11. The presence of HA did not have any influence on As(III) removal even at HA concentration up to 20 mg/L. Moreover, the addition of H2O2 promoted As(III) removal by P/Fe-CB0.5 and P/Fe-CB-3.0. The results suggested that P/Fe-CB could effectively remove arsenic in the absence or presence of HA. 3.6. As(III) adsorption mechanism 3.6.1. Effect of ionic strength The increase of ionic strength is known to significantly influence outer-sphere interactions between solute and adsorbent surface but not affect inner-sphere complexation [40]. The solution ionic strength was adjusted by the addition of NaCl. As shown in Fig. S12, the ionic strength showed a negligible effect on As(III) removal in the absence of H2O2, indicating that the As(III) adsorption onto the P/Fe-CB followed the inner-sphere complex mechanism. In contrast, the presence of NaCl only slightly suppressed the As(III) removal kinetics when H2O2 was added, and the inhibition was negligible at the end of the reaction. Thereby, the addition of H2O2 did not significantly change the arsenic adsorption mechanism. 3.6.2. FTIR analysis To further elucidate the As(III) adsorption mechanism, the FTIR characterizations of P/Fe-CB-3.0 before and after As(III) adsorption were performed (Fig. 6a). A new peak at 785 cm−1 appeared after As (III) adsorption. At pH 7, the dominant arsenite species in aqueous solution was As(OH)3 which had no IR bands in the 750-800 cm−1 region [41]. Therefore, the 785 cm−1should result from the formation of an inner-sphere surface complex. In the presence of H2O2, a new peak at 835 cm−1 appeared, which was ascribed to the As-O stretching in AsO-Fe linkage between As(V) and ferrihydrite [42]. This result indicated that: (1) the As(III) adsorbed on P/Fe-CB-3.0 was eventually oxidized to As(V) when H2O2 was added, and (2) both As(III) and As(V) adsorption on P/Fe-CB-3.0 followed the inner-sphere complex mechanism.
3.7. Performances of P/Fe-CB in different water matrices Natural water matrices are considerably more complex than deionized water because they contain various inorganic ions and organic matters. To better assess the feasibility of P/Fe-CB to As(III) removal under environmentally relevant conditions, three types of As(III)-spiked natural water matrices (groundwater, river water and lake water) were used. The water quality parameters are shown in Table S5. Fig. 7 showed the As(III) removal by P/Fe-CB-3.0 in the three water matrices. In all cases, the equilibrium As(III) concentration dropped from 200 μg/ L to below 10 μg/L within 360 min (Fig. 7a). When 0.1 mM H2O2 was added, the process could be shortened to 240 min (Fig. 7b). Similar results were obtained for P/Fe-CB-0.5 (Fig. S22). Thus, 1 L of As(III)contaminated natural water (200 μg/L) could be treated to meet the WHO standard for arsenic in drinking water using only 1 g of P/Fe-CB.
3.6.3. XPS analysis XPS measurements were performed to determine the surface chemistry of the adsorbents. As shown in Fig. 6b, new As3d as well as AsLMM peaks appeared after reaction with As(III) or As(III)/H2O2, suggesting that the arsenic was successfully bonded to P/Fe-CB-3.0. To reveal the arsenic species on P/Fe-CB-3.0, the As3d XPS spectra were analyzed (Fig. 6c). In the absence of H2O2, almost all the arsenic adsorbed on P/Fe-CB-3.0 was in the form of As(III). Comparatively, it was completely transformed to As(V) when H2O2 was added, agreeing well 102
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Fig. 6. FT-IR (a) and XPS (b) spectra of the P/Fe-CB-3.0 before and after As(III) or As(III)/H2O2 adsorption, (c) As3d XPS spectra, and (d) O1s XPS spectra.
3.8. Adsorbent regeneration and reuse
synthetic As(III)-contaminated groundwater containing As(III) and competing anions (CO32−, SO42−, SiO32−, PO43-). Their breakthrough curve is depicted in Fig. 8b. The P/Fe-CB-3.0 column generated about 3000 BV (27.12 L) effluent before the breakthrough point (10 μg/L) occurred. Additionally, the total Fe concentration leaching from the column was below 0.05 mg/L. The column arsenite adsorption density (qcolumn) at the breakthrough point was 1.19 mg/g, which was lower than the adsorption capacity from batch tests. This was due to the fact that the low initial As(III) concentration and very short contact time (4.5 min) of column operation made less As(III) be captured by P/FeCB-3.0. Thus, P/Fe-CB-3.0 was effective for As(III) removal from solution using bath batch or column operation.
Recyclability a crucial aspect to evaluate adsorbent applicability. Fig. 8a showed the successive adsorption − desorption cyclic runs of As (III) by P/Fe-CB-0.5 and P/Fe-CB-3.0. After each adsorption cycle, the adsorbent was regenerated with 0.1 M NaOH. In the absence of H2O2, the removal efficiency of As(III) reduced slightly as the regeneration cycle increased. After recycling four times, the removal efficiency of As (III) still maintained above 90%. More satisfactory results were achieved when 0.5 mM H2O2 was added and the removal efficiency was maintained at nearly 100% throughout five consecutive cycles. 3.9. Column operation of P/Fe-CB Fixed-bed adsorption was carried out to remove arsenite from
Fig. 7. As(III) removal in natural water samples by P/Fe-CB-3.0 (a) and in the presence of H2O2 (b). C0As(III) = 0.2 mg/L, adsorbent dose = 1 g/L, pH = 7, H2O2 = 0.1 mM (if added), and T = 25℃. 103
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Fig. 8. (a) As(III) removal in 5 consecutive cycles for P/Fe-CB, and (b) As(III) removal from a synthetic groundwater by using separate columns packed with P/Fe-CB3.0. C0As(III) = 0.975 mg/L for (a), adsorbent dose = 1 g/L, pH = 7, t = 12 h, and T = 25℃.
4. Conclusions
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Novel porous Fe-chitosan beads (P/Fe-CB) were prepared by the freeze-casting technique. The P/Fe-CB possessed unique pore structure and excellent acid resistance. The maximal adsorption capacities of P/ Fe-CB for As(III) was 52.7 mg/g at pH 7.0, which was higher than that for most of the reported granular adsorbents. Coexisting sulfate, carbonate, silicate and humic acid had no significant effect on As(III) removal, especially when H2O2 was added. The P/Fe-CB could be readily separated and regenerated, and the adsorption efficiency maintained above 90% throughout five consecutive cycles. For most situation, the addition of H2O2 accelerated the As(III) removal. In the presence of high concentration of phosphate (50 mg/L) or under high pH (11) condition, the addition of H2O2 would inhibit the As(III) removal. Moreover, it could effectively remove As(III) from natural water matrices. The fixed-bed working capacity was about 3000 BV. Therefore, the P/Fe-CB could be employed as a promising adsorbent for the As(III) removal from water. Acknowledgments This work was supported by the National Natural Science Foundation of China (51778218 and 51478171) and Technology Innovation Plan of Hunan Province (2017SK2420). Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.jhazmat.2019.03.070. References [1] D.K. Nordstrom, Worldwide occurrences of arsenic in ground water, Science 296 (2002) 2143–2145, https://doi.org/10.1126/science.1072375. [2] L. Winkel, M. Berg, M. Amini, S.J. Hug, C.A. Johnson, Predicting groundwater arsenic contamination in Southeast Asia from surface parameters, Nat. Geosci. 1 (2008) 536–542, https://doi.org/10.1038/ngeo254. [3] M. Argos, T. Kalra, P.J. Rathouz, Y. Chen, B. Pierce, F. Parvez, T. Islam, A. Ahmed, M. Rakibuz-Zaman, R. Hasan, Arsenic exposure from drinking water, and all-cause and chronic-disease mortalities in Bangladesh (HEALS): a prospective cohort study, Lancet 376 (2010) 252–258, https://doi.org/10.1016/S0140-6736(10)60481-3. [4] P. Smedley, D. Kinniburgh, A review of the source, behaviour and distribution of arsenic in natural waters, Appl. Geochem. 17 (2002) 517–568, https://doi.org/10. 1016/S0883-2927(02)00018-5. [5] C. Su, Environmental implications and applications of engineered nanoscale magnetite and its hybrid nanocomposites: a review of recent literature, J. Hazard. Mater. 322 (2017) 48–84, https://doi.org/10.1016/j.jhazmat.2016.06.060. [6] S. Dixit, J.G. Hering, Comparison of arsenic (V) and arsenic (III) sorption onto iron oxide minerals: implications for arsenic mobility, Environ. Sci. Technol. 37 (2003) 4182–4189, https://doi.org/10.1021/es030309t. [7] X. Zhang, C. Cheng, J. Qian, Z. Lu, S. Pan, B. Pan, Highly efficient water decontamination by using sub-10 nm FeOOH confined within millimeter-sized mesoporous polystyrene beads, Environ. Sci. Technol. 51 (2017) 9210–9218, https://doi.org/
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