The removal of chromium (VI) and lead (II) from groundwater using sepiolite-supported nanoscale zero-valent iron (S-NZVI)

Chemosphere 138 (2015) 726–734

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The removal of chromium (VI) and lead (II) from groundwater using sepiolite-supported nanoscale zero-valent iron (S-NZVI) Rongbing Fu a,⇑, Yingpin Yang b, Zhen Xu a, Xian Zhang b, Xiaopin Guo a, Dongsu Bi b a b

Shanghai Academy of Environmental Sciences, Shanghai 200233, China Department of Chemical Engineering, Shanghai Institute of Technology, Shanghai 201418, China

h i g h l i g h t s  NZVI was immobilized on sepiolite to overcome agglomeration.  Heavy metals were rapidly and extensively remediated by sepiolite-supported NZVI.  The remediation of metals occurred in a stepwise manner (adsorption–reduction).  Sepiolite-supported NZVI could be applied to in situ heavy metal remediation.

a r t i c l e

i n f o

Article history: Received 7 February 2015 Received in revised form 20 June 2015 Accepted 19 July 2015

Keywords: Sepiolite Chromium (VI) Lead (II) Nanoscale zero-valent iron Groundwater

a b s t r a c t In this study, the synthesis and characterization of sepiolite-supported nanoscale zero-valent iron particles (S-NZVI) was investigated for the adsorption/reduction of Cr(VI) and Pb(II) ions. Nanoscale zero-valent iron (NZVI) supported on sepiolite was successfully used to remove Cr(VI) and Pb(II) from groundwater with high efficiency. The removal mechanism was proposed as a two-step interaction including both the physical adsorption of Cr(VI) and Pb(II) on the surface or inner layers of the sepiolite-supported NZVI particles and the subsequent reduction of Cr(VI) to Cr(III) and Pb(II) to Pb(0) by NZVI. The immobilization of the NZVI particles on the surface of sepiolite could help to overcome the disadvantage of NZVI particles, which have strong tendency to agglomerate into larger particles, resulting in an adverse effect on both the effective surface area and reaction performance. The techniques of XRD, XPS, BET, Zeta potential, and TEM were used to characterize the S-NZVI and interaction between S-NZVI and heavy metals. The appropriate S-NZVI dosage was 1.6 g L1. The removal efficiency of Cr(VI) and Pb(II) by S-NZVI was not affected to any considerable extent by the presence of co-existing ions, such 2 2+ and HCO as H2PO 4 , SiO3 , Ca 3 . The Cr(VI) and Pb(II) removal kinetics followed a pseudo-first-order rate expression, and both Langmuir isotherm model and Freundlich isotherm model were proposed. The results suggested that supporting NZVI on sepiolite had the potential to become a promising technique for in situ heavy metal-contaminated groundwater remediation. Ó 2015 Elsevier Ltd. All rights reserved.

1. Introduction Land and water pollution by heavy metals is a worldwide issue. Chromium (VI) and lead (II) are two common toxic metals found in both soil and groundwater. Cr(VI) is an industrial contaminant in both soil and groundwater from such industries as leather tanning, electroplating, metal processing, film processing, and mining of chromium ores (Sarin et al., 2006). Chromium (VI) engenders environmental and health problems because of its potent toxicity, mutagenicity and carcinogenicity. It is highly soluble in water ⇑ Corresponding author. E-mail address: [email protected] (R. Fu). http://dx.doi.org/10.1016/j.chemosphere.2015.07.051 0045-6535/Ó 2015 Elsevier Ltd. All rights reserved.

and toxic to many different aquatic organisms even at low concentrations (Mohan and Pittman Jr., 2006). Therefore, chromium (VI) is regarded as the priority controlled pollutant in many countries (Hizal and Apak, 2006). Pb(II) is widely used in many industrial applications, such as storage battery manufacturing, painting pigment, fuels, photographic materials, coatings, and the automotive and aeronautical industries (Lalhruaitluanga et al., 2010). This heavy metal is a highly toxic and cumulative poison that accumulates mainly in the bones, brain, kidneys and muscles. Lead poisoning in humans causes severe damage to the kidneys, nervous system, reproductive system, liver and brain (S ß ölener et al., 2008). Therefore, it is very important to remove heavy metals before disposal (Liu et al., 2014).

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Nanoscale zero-valent iron (NZVI) has been demonstrated to be a good alternative for the remediation of a wide variety of organic and inorganic pollutants in soil and groundwater, showing a high efficiency and low economic and environmental costs (Cundy et al., 2008). The NZVI has high reactivity and surface energy due to its large SSA (Sun et al., 2006), more reactive surface sites, and high reaction rate (Li et al., 2006). Nevertheless, there are still some technical challenges in NZVI applications that affect the use of this material in in-situ remediation. For instance, the aggregation of NZVI limits its mobility, dispersivity, durability, and mechanical strength, and the oxidization of NZVI can significantly decrease its reactivity. To address these issues, technologies have attempted to develop material (kaolin (Zhang et al., 2010a), bentonite (Shi et al., 2011a), montmorillonite (Zhang et al., 2013), zeolite (Kim et al., 2013), and rectorite (Luo et al., 2013) ), attapulgite (Guixiang Quan et al., 2014) to support NZVI to decrease aggregation and improve its dispersivity. In addition, (Fei et al., 2012) had reported the use of sepiolite as support of NZVI to degrade bromamine acid. Sepiolite is a clay mineral with a wide range of applications based on its physicochemical properties and especially surface properties. It is chemically stable, inexpensive and an ideal support material for NZVI (Suarez and García-Romero, 2012). It is a hydrous magnesium silicate (Si12O30Mg8(OH)4(H2O)48H2O) characterized by its fibrous morphology that is due to its crystalline structure (Marjanovic´ et al., 2013). It presents a structure of needle-like particles and has talc-like layers that consist of two layers of tetrahedral silica and a central octahedral magnesium layer (Lazarevic´ et al., 2012). Due to its sorptive, rheological and catalytic properties, sepiolite is widely used in a variety of industrial applications (Lazarevic´ et al., 2012). However, its relatively low surface acidity, narrow channels, small surface area, and poor thermal stability seriously limit applications of the natural sepiolite (Liang, 2008). Therefore, it is essential to modify the sepiolite before synthesizing sepiolite–supported materials. In this study, we investigated the feasibility of using sepiolite-supported nanoscale zero-valent iron particles (S-NZVI) to remove Cr(VI) and Pb(II) from groundwater. We developed a two-step reduction technique to successfully synthesize S-NZVI particles at normal pressure and temperature. The synthesized material before and after reaction was characterized by X-ray diffraction (XRD), a BET-N2 adsorption technique, transmission electron microscopy (TEM), Zeta potential analysis and X-ray photoelectron spectroscopy (XPS). The reaction mechanism was studied and proposed. In addition, the effects of pH, S-NZVI dosage, initial heavy metal concentration and coexisting ions on Cr(VI) and Pb(II) removal were also evaluated.

dispersion of 40 g of sepiolite in 200 mL of a 4 M HCl solution and stirring the suspension for 48 h at room temperature. The solid was then separated from the solution by centrifugation and washed with deionized water until it was Cl ion free. The obtained acid-activated sample was dried at 110 °C for 2 h and calcinated at 300 °C for 4 h and then sieved through a 160 mesh screen prior to use. The S-NZVI with a sepiolite/iron mass ratio of 9:1 was prepared by the following method (Fig. 1): Asp (7.5 g) was added to 150 mL of a Fe3+ solution (0.1 M) and stirred overnight using a mechanical stirrer at room temperature. One-hundred fifty millilitres of 0.4 M NaBH4 was added drop-wise into the suspension under vigorous agitation and a nitrogen atmosphere. The sepiolite particles in the solution turned black immediately and the mixture was sequentially stirred for 0.5 h. The solid was separated from the suspension by vacuum filtration and washed several times with deionized water followed by absolute ethyl alcohol. Finally, S-NZVI was dried under vacuum at 60 °C overnight. The deoxygenated water used in the experiment was boiled for 0.5 h and then purged with nitrogen for another 0.5 h to remove dissolved oxygen. Conventionally synthesized NZVI was prepared by the Fe3+ reduction method. Briefly, 1.6 M freshly prepared NaBH4 aqueous solution was carefully added drop-wise to the 0.4 M FeCl36H2O solution at room temperature with continuously stirring under a gentle stream of nitrogen (Shi et al., 2011a).

2. Materials and methods

2.4. Batch experiments

2.1. Materials and chemicals

The concentration of Pb(II) in solution was determined using an inductively coupled plasma emission spectrometer ICP (Optima 7000DV, Perkin–Elmer, USA). The concentration of Cr(VI) was measured by the 1,5-diphenylcarbazide method at a wavelength of 540 nm on a UV–VIS spectrophotometer (TU-1810, Beijing General Instrument Co., Ltd. China). All of the experiments were conducted in duplicate. Each batch reaction took place in a 50 mL centrifuge tube containing a given amount of S-NZVI and 25 mL of a solution with a certain concentration of Cr(VI) and Pb(II) solution. These tubes were placed in a steam-bath vibrator (Model ZD-85, China) at a speed of 200 rpm at 28 °C. After a given time interval, several vials were taken out of the vibrator. The initial pH values of the suspensions were adjusted using 0.1 M HCl or NaOH. The particles and an aqueous phase were separated by centrifugation at 8000 rpm for 3 min. The supernatant samples were filtered through a 0.22-lm

Iron (III) chloride hexahydrate (FeCl36H2O), sodium borohydride (NaBH4), sodium hydroxide (NaOH), hydrochloric acid (HCl), sulphuric acid (H2SO4), phosphoric acid (H3PO4), nitric acid (HNO3), calcium chloride (CaCl2), sodium dihydrogen phosphate (NaH2PO4), sodium bicarbonate (NaHCO3), sodium metasilicate (Na2SiO3), potassium dichromate (K2Cr2O7), and lead nitrate (Pb(NO3)2) were purchased from the Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China). All chemicals were of analytical reagent grade and used without further purification. 2.2. Preparation of NZVI and S-NZVI Sepiolite was purchased from Nanyang Boxing Mining Co., Ltd. (Hebei, PRC). Acid-activated sepiolite (Asp) was prepared by the

2.3. Characterizations The morphological analysis of NZVI, S-NZVI and sepiolite was performed using a transmission electron microscope (TEM, TECNAI G2 F30, America FEI, USA). The X-ray diffraction (XRD) patterns of sepiolite, S-NZVI, and NZVI were determined using an X-ray diffractometer (PANalytical-X’Pert PW3040/60, the Netherlands) with a high-power Cu Ka radioactive source (40 kV, 40 mA), and X-ray photoelectron spectroscopy (XPS, PHI5000C ESCA System) analysis was carried out to characterize the S-NZVI particles before and after exposure to Cr(VI) and Pb(II). All of the samples were scanned from 3° to 90° (2h) and The 2h-scanning rate was 0.2089°s1. The specific surface areas (SSA) of sepiolite, Asp, NZVI, and S-NZVI were determined by the Brunauer–Emmet–Tell er (BET) technique with an accelerated surface area and porosimetry analyser (Micromeritics’ ASAP 2020 Instrument Crop., USA). The zeta potentials were measured by a Zetasizer Nano zeta potential analyser (Malvern Instruments, UK, motion range: ±10 lm/vs conductivity Range: 0–200 ms/cm, minimum volume: 0.75 mL, temperature: 2–90 °C).

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Fig. 1. Schematic illustration of the immobilization of nanoscale zero-valent iron particles on sepiolite (a). A proposed mechanism for Cr (VI) and Pb(II) remediation by S-NZVI (b).

membrane filter. The comparison tests of acid-activated sepiolite, NZVI and S-NZVI were also conducted. Water to prepare the solution was boiled for 0.5 h and subsequently purged with N2 for 0.5 h to remove dissolved oxygen. 2.5. Adsorption isotherms Adsorption equilibrium of S-NZVI with Cr(VI) and Pb(II) can be described by Langmuir and Freundlich. The Langmuir isotherms expressed in Eq. (1) and the Freundlich isotherms is given by Eq. (2).

Ct 1 Ct ¼ þ Q t Q m kL Q m

ð1Þ

lgCt þlgkF n

ð2Þ

lgQ t ¼

where Qt is the amount of adsorbate per mass of adsorbent at equilibrium (mg g1), Ct is the equilibrium concentration of adsorbate in aqueous solution (mg L1), Qm is the monolayer adsorption capacity at equilibrium (mg g1), kL is the Langmuir equilibrium constant (L mg1), kF is a Freundlich constant (index of adsorption capacity), and n is the Freundlich constant (index of adsorption intensity or surface heterogeneity). 3. Results and discussion 3.1. Characterization and interaction mechanism The removal mechanism of Cr(VI) and Pb(II) by S-NZVI is outlined in Fig. 1. A two-step interaction mechanism is proposed (Li et al., 2012). First, Cr(VI) and Pb(II) in aqueous solution were absorbed on the S-NZVI surface. Second, Cr(VI) and Pb(II) were reduced by Fe0 and co-precipitated or precipitated on the surface of S-NZVI. Cr(VI) was reduced to the less toxic Cr(III) by S-NZVI and immobilized by precipitation as Cr(OH)3 or by incorporation into the iron oxide/iron hydroxide shell forming alloy-like Cr3+– Fe3+ hydroxides (Ponder et al., 2000; Shi et al., 2011a), but the formation of these Cr3+–Fe3+ hydroxides may inhibit electron transfer from the Fe0 to the surface, thus stopping the redox process and causing a decrease of Cr(VI) removal, especially at high Cr(VI) concentrations (Hu et al., 2010; O’Carroll et al., 2013). The surface morphology of the sepiolite, NZVI and S-NZVI that were analysed by TEM is presented in Fig. 2. It was observed that

the conventionally synthesized NZVI particles (Fig. 2a) appeared to be spherical in the size range of 20–60 nm. They exist as prominent chain-like aggregates with lengths of 50 nm–2 lm because of their high surface energies and magnetic interactions. The NZVI particles (D = 10–50 nm) immobilized on sepiolite were clearly discrete and well dispersed on the sepiolite support that itself had a needle-like morphology (Fig. S1) without aggregation (Fig. 2b). Therefore, the nanoparticles of iron were in the form of spherical particles, which were very different from conventionally synthesized NZVI. Other materials, such as resin, starch, and zeolite, were also useful in enhancing the dispersibility of NZVI particles (Shu et al., 2010; Shi et al., 2011a; Kim et al., 2013; Mosaferi et al., 2014) because of a decrease in aggregation and an increase in mechanical strength. Fig. 2b and c show the changes in the morphology of S-NZVI before and after reaction. Before reaction (Fig. 2b), NZVI particles were distributed uniformly on the surface of sepiolite with an elemental composition of Fe, O, Si (Fig. 2e). However, after S-NZVI reacted with Cr(VI) and Pb(II), as shown in Fig. 2c and d, flocculent clusters with diffuse edges formed on the surface of reacted S-NZVI composites indicating the presence of elemental Pb and Cr and their oxides via redox reactions. This changed the morphology of S-NZVI (Xi et al., 2010). This could be confirmed by the elemental distribution of S-NZVI before and after reaction with Cr(VI) and Pb(II) using SEM-EDX (Fig. 2e, f and g). The EDX result showed that S-NZVI before reaction mainly consisted of Fe, O, and Si. The O and Si were from sepiolite. The Fe came from NZVI (Su et al., 2011). After reaction with Cr(VI) and Pb(II) (Fig. 2f and g), Cr and Pb were introduced, respectively. In addition, the Fe content in S-NZVI dropped from 63.88% to 56.82% and 42.71% (w:w), while the O content increased from 15.26% to 26.61% and 32.34% (w:w) compared with the blank S-NZVI. This could be explained by the oxidation of Fe0 as a reductant to form the Fe2+ and Fe3+ iron oxide or hydroxide in the S-NZVI (Su et al., 2011). The XRD patterns of sepiolite, Asp and S-NZVI before and after reaction with Cr(VI) and Pb(II) are illustrated in Fig. 3. The typical peak of MgO (2h at 43.277°, d = 0.20889 nm) became too weak to be visible in Asp, and other typical peaks of sepiolite remained, which indicated that Mg2+ was replaced by H+ under acidic conditions. Before reaction, a diffraction peak at 44.67° (d = 0.22077 nm) (Fig. 3c) corresponding to a formation of zero-valent iron (Fe0) could be observed, indicating that Fe0 nanoparticles were incorporated with Asp. After activation of sepiolite particles by HCl solution their crystalline structure was significantly decreased and

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Fig. 3. The X-ray diffractograms of (a) sepiolite; (b) Asp; (c) S-NZVI; (d) S-NZVI after reaction with Pb(II) and (e) S-NZVI after reaction with Cr(VI).

Fig. 2. The TEM image of NZVI (a), S-NZVI (b), S-NZVI after reaction with Cr(VI) (c) and (d) the TEM image of S-NZVI after reaction with Pb(II); the EDX survey for SNZVI (e), S-NZVI after reaction with Cr(VI) (f) and (g) EDX survey for S-NZVI after reaction with Pb(II).

did not affect the performance of NZVI particles. Both the EDX and the XRD analysis demonstrate that the metallic iron products have been obtained and the iron nanostructures are very stable under ambient conditions (Tong et al., 2008)_ENREF_21. However, after S-NZVI reacted with Cr(VI) (Fig. 3d) and Pb(II) (Fig. 3e), the typical peak of Fe0 (2h at 44.67°) became too weak to be visible, while the peaks of Fe2O3 (2h at 35.6° and 41.2°), FeO(OH) (2h at 29.5°), Fe3O4 (2h at 35.4°/56.6°/62.2°), Cr2FeO4, Cr2O3 (2h at 33.4°, d = 0.26806 nm and 42.699°, d = 0.21158 nm) and Pb0, PbO2, PbOxH2O (2h at 42.537°, d = 0.21235 nm and 32.723°, d = 0.27244 nm) appeared more obvious. These observations demonstrated the adsorption or co-precipitation of the products on the surface of the S-NZVI particles. These products resulted from the redox reaction between Fe0, Fe2+ or H+ and Cr(VI) or Pb(II) where the supported NZVI acted as a reductant in aqueous solution.

The SSA of sepiolite, Asp, NZVI and S-NZVI were measured by the BET-N2-adsorption–desorption method. Certain amounts of structural Mg2+ ions were removed (Lazarevic´ et al., 2007), and the SSA and total pore volume increased to 230.37 m2 g1 and 0.29 cm3 g1 (Table S1) after an acidic treatment of sepiolite. Previous studies also confirmed that acidic treatment increased the surface area of sepiolite. Solids were obtained that had high porosity and a high number of acidic centres (Lazarevic´ et al., 2007). When NZVI was loaded onto sepiolite, a mean SSA of 141.42 m2 g1 was obtained, which was much higher than that of NZVI (47.74 m2 g1) and other results (Table S2). Compared to NZVI, the increased SSA of S-NZVI was due to the non-aggregation of NZVI particles (Kim et al., 2013). The TEM image of S-NZVI (Fig. 2c) showed that the NZVI particles were dispersed onto the surface of sepiolite, which indicated that sepiolite played a role in dispersing and stabilizing NZVI and that the supported nanoparticles might exhibit higher reaction rates with contaminants compared to unsupported particles. This was confirmed by previous studies (Fei et al., 2012). Arshadi et al. synthesized S-NaOH-NZVI with a lower SSA but higher reaction activity than that of other synthesized iron nanoparticles (Arshadi et al., 2014). The supported NZVI was immobilized on the active sites of sepiolite, which would decrease the oxidation of the NZVI surface (Arshadi et al., 2014). Generally, particles with a zeta potential from 15 mV to +15 mV are expected to be unstable from electrostatic considerations (Hiemenz, 1977). The zeta potentials of sepiolite, Asp, NZVI and S-NZVI were 26.1 mV, 22.7 mV, +14 mV and 44 mV under neutral condition (Fig. S2), respectively, indicating that the S-NZVI was most stable among these materials. After loading NZVI on Asp, the zeta potential changed to 44.0 mV. However, the zeta potential of S-NZVI after reaction with Cr(VI) (Fig. S2e) and Pb(II) (Fig. S2f) were +0.76 mV and +7.9 mV, respectively, because the surface charge of the S-NZVI was dominated largely by Cr(III) and Pb(II) owing to the negative surface charges of S-NZVI which attracted the cations (Xu et al., 2013). The results of the XPS characterization are shown in Fig. 4. The new peaks at the binding energy (BE) of 580 eV and 140 eV appearing after the reduction of the chromium and lead (Fig. 4a) were not obvious, which might be due to a part of heavy metal adsorbed by sepiolite entering the internal channels of sepiolite. The intense signal C 1s might be due to the fiber of filter paper or carbon dioxide contamination from the air, water or other containers during the sample preparation and transfer (Li et al., 2008).

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Fig. 4. Typical wide scan XPS spectra for S-NZVI before and after Cr(VI) and Pb(II) reactions (a), high-resolution XPS survey of iron 2p (b), oxygen 1s (c), chromium 2p (d) and lead 4f (e).

Detailed XPS surveys of the region of Fe 2p, O 1s, Cr 2p and Pb 4f are presented in Fig. 4b, c and d. In Fig. 4b, the photoelectron peaks approximately 711–724 eV corresponded to the binding energies of 2p3/2 and 2p1/2 of Fe2+ and Fe3+. It is very difficult to detect the presence of Fe° and it seems to be background noise. The photoelectron peak for the oxygen 1s region resolved into three curves with peaks of approximately 529, 531 and 532 eV, which represented the binding energies of oxygen in OH, O2 and chemically or physically adsorbed water (H2O) (Li and Zhang, 2007; Lv et al., 2012), respectively. In the Fig. 4c, there was a significant shift of peak positions before and after the S-NZVI reacted with Cr(VI) and Pb(II). The Cr(VI) and Pb(II) sorption was accompanied by a change in oxygen binding, providing evidence that the oxygen-containing functional groups on the surface of the S-NZVI take part in the sorption of Cr(VI) and Pb(II) (Xu et al., 2008). The oxygen peak shifted obviously due to the pellets of Cr(OH)3 and Pb(OH)2 begin to occur and cover the S-NZVI surface. In Fig. 4d and e, the photoelectron peaks for chromium 2p3/2, 2p1/2, and lead 4f5/2, 4f7/2 centres at 580.1 eV, 589.8 eV and 146.1 eV, 141.3 eV after 60 min of treatment, respectively. In Fig. 4d, the Cr 2p3/2 region could be decomposed into two peaks at 576.6 and 577.6 eV, implying that Cr(III) and Cr(VI) coexisted on these extracted particles. The binding energies of the Cr 2p peaks are in good agreement with those reported for FeCr2O4 and Cr2O3 or Cr(OH)3 (Nahuel Montesinos et al., 2014). This result demonstrated that both adsorption of Cr(VI) and reduction of Cr(VI) on the surface of the S-NZVI occurred simultaneously (Geng et al., 2009). As shown in Fig. 4e, Pb 4f dominated with two peaks at 136 eV and 141 eV, which can be assigned to Pb(0) and Pb(II), respectively, according to previous studies (Huang et al., 2013). In this study,

Pb(0) was observed on the particle surface with a peak at 136.6 eV, and these values are only a little higher than those obtained in a previous study (Xi et al., 2010). It can be concluded that the presence of a portion of Cr(III) and Pb(0) on the surface of the S-NZVI can be solely attributed to the reduction by Fe0 but the generation of Fe(III) can be caused by the synergetic oxidation of Cr(VI), Pb(II) and by a simultaneous redox reaction between water and O2 in the experimental system (Geng et al., 2009). 3.2. Effect of pH, S-NZVI dosage, initial concentration and co-existing ions The pH value, S-NZVI dosage, initial concentration and co-existing ions are important factors affecting the removal efficiency of Cr(VI) and Pb(II) (Fig. 5). The pH value of natural groundwater commonly ranges from 5.0 to 9.0. We chose 4.0–9.0 to assess the pH effect on the removal of Cr(VI) and Pb(II) by S-NZVI, and the result is shown in Fig. 5a and b. The removal efficiency of Cr (VI) decreased with the increase in pH (Fig. 5a). From pH 4.0 to 9.0, the removal rate decreased from 98.5% to 30.7%, which indicated that the removal of Cr(VI) by NZVI was strongly pH-dependent, and an acidic environment favoured Cr(VI) removal by the S-NZVI. Fig. 5b showed the effect of pH on Pb(II) removal by the S-NZVI. It was observed that by increasing the pH from 4.0 to 9.0, the Pb(II) sorption decreased from 99.9% to 80.3%. With an increase of pH from 4.0 to 9.0, the pH value did not significantly affect the Pb(II) removal. Both the removal of Cr(VI) and Pb(II) decreased with an increase in pH. The kobs (Table 1) of Cr(VI) and Pb(II) indicated that the highest kobs of Cr(VI) and Pb(II) were 0.754 min1 and 1.26 min1,

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Fig. 5. Effect of different conditions on Cr(VI) and Pb(II) removal by S-NZVI; (a), (b) effect of pH; (c), (d) effect of S-NZVI dosage; (e), (f) effect of the initial concentration; (g), (h) effects of co-existing ions.

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Table 1 Pseudo-first-order adsorption kinetics constants for Cr(VI), Pb(II) removal by S-NZVI composites under different pH, S-NZVI dosage and initial concentration of Cr(VI) and Pb(II) (S-NZVI = 1.6 g L1, t = 60 min). pH

Cr Pb

kobs (min1) R2 kobs (min1) R2

4.0

5.0

6.0

7.0

8.0

9.0

0.652 0.93 0.11 0.98

0.711 0.98 0.165 0.99

0.754 0.98 1.26 0.99

0.022 0.98 0.026 0.97

0.005 0.97 0.015 0.94

0.004 0.93 0.006 0.98

S-NZVI Dosage (g L1)

Cr Pb

kobs (min1) R2 kobs (min1) R2

0.05

0.1

0.4

0.8

1.6

3.2

0.002 0.90 0.004 0.94

0.009 0.91 0.008 0.95

0.036 0.91 0.013 0.98

0.054 0.91 0.028 0.99

0.52 0.94 0.076 0.98

0.56 0.92 0.082 0.96

Initial concentration (mg L1)

Cr Pb

1

kobs (min ) R2 kobs (min1) R2

20

40

50

60

70

80

100

120

0.082 0.97 1.82 0.93

0.29 0.86 1.91 0.99

0.094 0.98 1.32 0.99

0.0042 091 1.27 0.99

0.012 0.99 1.05 0.99

0.0048 0.90 0.014 0.99

0.0035 0.95 0.016 0.99

0.006 0.96 0.008 0.99

respectively, at pH 6.0. In Fe0 treatment systems, the removal mechanisms of Cr(VI) and Pb(II) are generally believed to involve the adsorption of Cr(VI) and Pb(II) on the iron surface where electron transfer takes place, and then, Cr(VI) is reduced to Cr(III) and Pb(II) is reduced to Pb(0) with the oxidation of Fe0 to Fe3+ or Fe2+ under acidic conditions (O’Carroll et al., 2013). At the same time, precipitates formed by iron hydroxides alone or co-precipitates of chromium (or lead) have been observed by other researchers at higher pH values (Powell et al., 1995). In the range of pH 4.0–6.0, the removal of Pb and Cr remains constant and reaches a maximum. The removal of Pb(II) and Cr(VI) is mainly accomplished by sorption reactions (Xu et al., 2008). At high pH values (pH 7.0–9.0), the adsorption of Cr(VI) was decreased due to electrostatic repulsion caused by the negative NZVI surface charge above the pH of 7.8 (O’Carroll et al., 2013). The higher the valence of adsorbed anions, the more negative the surface becomes, hence the inhibition of the further adsorption of anions (Hu et al., 2009). The decrease of Pb(II) sorption on S-NZVI was attributed to the increase in the hydroxyl(–OH) concentration, which leads to a negatively charged S-NZVI surface and caused precipitation of Pb(OH)2 on the S-NZVI surface as well (Zhang et al., 2010a). In addition, the Fe (oxy) hydroxide coating on the NZVI surface at a pH above 7.0 would also occupy the active sites on the NZVI surfaces and decrease reactivity. The impact of the S-NZVI dosage (0.05–3.2 g L1) on the removal reaction was determined (Fig. 5c and d). The removals of Cr(VI) and Pb(II) of 98.8% and 97.31% were obtained within 10 min when the S-NZVI addition was 3.2 g L1. As the S-NZVI dosage increased from 0.05 g L1 to 3.2 g L1, the removal efficiency of Cr(VI) increased significantly from 45.1% to 99.2% and that of Pb(II) increased from 56.2% to 99.9%, respectively. Meanwhile, kobs (Table 1) increased as the S-NZVI dosage increased. In light of the previously published data (Shu et al., 2007; Chen et al., 2011; Shih et al., 2011), the removal efficiency or rate constants increase with the increase of material addition. According to the results, the highest removal capacity for Cr(VI) and Pb(II) were 609.5 mg Cr/g Fe and 756.7 mg Pb/g Fe which was much higher than a few reports (25–100 mg Cr/g Fe (Shi et al., 2011b), 50–180 mg Cr/g Fe (Li et al., 2008), 344.8 mg Pb/g Fe (the highest removal capacity) (Kim et al., 2013), 241.75 mg Pb/g Fe (Zhang et al., 2010b)). However, as the dosage of S-NZVI increased from 1.6 g L1 to 3.2 g L1, the kobs of Cr(VI) and Pb(II)

was not affected significantly (Cr(VI) 0.52–0.56 min1; Pb(II) 0.076–0.082 min1). Therefore, it is not necessary to use a much higher S-NZVI dosage. A dosage of 1.6 g L1 S-NZVI was chosen as the optimal dosage. The effect of the initial concentration of Cr(VI) and Pb(II) on heavy metal (Cr(VI), Pb(II)) removal is shown in Fig. 5e and f. It can be observed that the removal rate of Cr(VI) and Pb(II) decreased with an increase in the initial concentrations of Cr(VI) and Pb(II). The pseudo-first-order model provided the best fit based on the experimental data as shown in Table 1. The removal kobs of Cr(VI) and Pb(II) decreased with the increase of the initial concentration of Cr(VI) and Pb(II). It has been reported that both Cr(VI) and Pb(II) are oxidants and passivators of Fe0; as more Cr(VI) or Pb(II) approached Fe0, more Fe0 would be oxidized and would lose its activity leading to a decrease in the removal rate. A precipitate forming on the surface of the NZVI would reduce the electron transfer from the NZVI to Cr(VI) and Pb(II) and would accordingly retard the reduction of Cr(VI) or Pb(II). Furthermore, for a fixed adsorbent dosage, the total available active sites are limited, thus leading to a decrease in the percentage removal of adsorbate corresponding to an increased initial adsorbate concentration. The higher removal efficiency of the S-NZVI at low Cr(VI) and Pb(II) initial concentrations could be related to the high ratio of initial mole numbers of Cr(VI) and Pb(II) to the available active sites on the surface area. Therefore, the Cr(VI) and Pb(II) removal is dependent on the initial concentration. There are various types of co-existing ions in the groundwater, such as calcium, bicarbonate, phosphate and silicate (Lv et al., 2013). As shown in Fig. 5g and h, the effect of the common co-existing ions on Cr(VI) and Pb(II) removal were investigated. From figure, it is evident that the Cr(VI) and Pb(II) removal efficiency of S-NZVI was not obviously affected by the presence of co-existing ions. However, the existence of HCO 3 did slightly affect Cr(VI) removal by the S-NZVI; the removal efficiency decreased from 99.80% to 88.18%. According to the principle of oxidation and mass balance, the reaction was conducive to the formation of Cr(III). However, the flocculent precipitate contributed to the formation of a precipitate and Cr(III) decreased in the reduction reaction when HCO 3 was in the system. The competition between HCO 3 and Cr(VI) for active sites on the S-NZVI affected the Cr(VI) removal; the HCO 3 could form inner-sphere surface complexes with iron (oxy)hydroxides and reduce the adsorption capacity

R. Fu et al. / Chemosphere 138 (2015) 726–734

(Giasuddin et al., 2007; Tanboonchuy et al., 2012) as well_ENREF_26. Moreover, the Fe2+ that is generated could react with HCO 3 to form FeCO3, and it would drop out of solution as a precipitate due to its limited solubility (Heuer and Stubbins, 1999). Besides, with the increasing of HCO 3 concentration, the hydrolyzation reaction of metal aggravated (Du et al., 2008) and the pH increased, which could inhibit electron transfer in the system. Therefore the removal efficiency was affected. 3.3. Comparison of Cr(VI) and Pb(II) removal by Asp, NZVI and S-NZVI The comparison of Cr(VI) and Pb(II) removal by sepiolite, acid activated sepiolite, NZVI and S-NZVI was show in Fig. S3. The removal efficiency of Cr(VI) and Pb(II) from 25 mL aqueous solution containing an initial concentration of 20 mg L1 using acid activated sepiolite (0.09 g), NZVI (0.01 g) and S-NZVI (0.1 g) after 1 h incubation at pH 6.0. In Fig. S3, S-NZVI displayed the highest Cr(VI) and Pb(II) removal efficiency in the solution. The removal efficiency of Cr(VI) and Pb(II) were 8.147% and 26.713% by acid activated sepiolite which was lower than that of Cr(VI) and Pb(II). This indicated that the removal of Cr(VI) and Pb(II) from solution using S-NZVI was much higher than that of acid activated sepiolite and NZVI. 3.4. Adsorption isotherm analysis The equilibrium isotherms for the adsorption of Cr(VI) and Pb(II) ions by the S-NZVI are presented in Table 2. It can be seen that the adsorption capacity increases with an increase in the equilibrium metal ion concentration for both metal ions in the solution. The equilibrium data are fitted by Langmuir and Freundlich adsorption isotherms. The isotherm parameters and related correlation coefficients (R2) were calculated from the slope. As displayed in Table 2, the results indicated that the adsorption isotherm data of both metals were well fitted to both the Langmuir isotherm model and the Freundlich model. Based on the Langmuir isotherms, the maximum adsorption capacity of the S-NZVI toward Cr(VI) and Pb(II) is 43.86 mg g1 (0.84 mmol g1) and 44.05 mg g1 (0.16 mmol g1), respectively. This is higher than the adsorption capacity of most reported adsorbents: activated alumina (Cr(VI): 25.57 mg g1) (Bhattacharya et al., 2008), iron oxide nanoparticles (Cr(VI): 34.1 mg g1; Pb(II) 36.0 mg g1) (Nassar, 2010; Kim et al., 2012), oxidized multiwalled carbon nanotubes (Pb(II): 9.92  103 mmol g1) (Xu et al., 2008), granular activated carbon (Cr(VI): 3.9 mg g1; Pb(II) 21.5 mg g1) (Rivera-Utrilla et al., 2003), montmorillonite (Pb(II): 33 mg g1) (Bhattacharyya and Gupta, 2007). The higher R2 of the Freundlich model implied a greater tendency of the heterogeneous surface of S-NZVI (Lv et al., 2012) for the removal of Cr(VI) and Pb(II). The low value of 1/n (<0.36) in the Freundlich isotherm suggested that any large change in the equilibrium concentration of chromium and lead would not result in a significant change in the amount of chromium and lead adsorbed by S-NZVI (Nguyen et al., 2010). The Freundlich constant n is found to be greater than 1 which indicates a favourable condition for adsorption (Badruddoza et al., 2013). Table 2 Langmuir and Freundlich isotherm parameters for Cr(VI) and Pb(II) removal by S-NZVI. Langmuir isotherm Cr(VI)

Pb(II)

kL (L mg1) Qm (mg g1) R2 kL (L mg1) Qm (mg g1) R2

Freundlich isotherm 0.052 43.86 0.9523 0.073 44.05 0.9504

kF 1/n R2 kF 1/n R2

2.61 0.31 0.9772 3.76 0.34 0.9575

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4. Conclusions In this study, S-NZVI was successfully prepared with a high capacity for and efficiency of removal of Cr(VI) and Pb(II) from groundwater. The NZVI particles with a diameter of 10–50 nm were distributed uniformly on the surface of sepiolite, demonstrating that modified sepiolite was effective in preventing NZVI particles from agglomerating. The removal rates of Cr(VI) and Pb(II) decreased with the increasing initial pH value and initial concentration of Cr(VI) and Pb(II). The removal efficiency of Cr(VI) and Pb(II) by S-NZVI was not affected to considerable extent by the presence of co-existing ions. Optimal removal occurred at pH 6.0 and a S-NZVI dosage of 1.6 g L1. The adsorption equilibrium data for both heavy metals were well fitted by both Langmuir and Freundlich isotherm models. The kinetic data were closely fitted to the pseudo-first-order model. The results of the XRD, XPS, BET, Zeta potential, and TEM techniques indicated that the removal mechanism was a two-step interaction, including both physical adsorption of Cr(VI) and Pb(II) on the surface or internal layers of S-NZVI particles and the subsequent reduction of Cr(VI) to Cr(III) and Pb(II) to Pb(0) by the S-NZVI. The results indicated that S-NZVI was effective in the removal of various heavy metals from groundwater, in terms of high efficiency, stability, and reactivity. Such a material can be a potential candidate for the in situ remediation of heavy metals. Acknowledgments This research was supported by the National Nature Science Foundation of China (Grant No. 41372262), the National Nature Science Foundation of China (Grant No. 41301344), the Nature Science Foundation of Shanghai (Grant No. 13ZR1435100), the Talent Development Fund Project of Shanghai (Grant No. 201325), and the Shanghai Environmental Protection Scientific Research Grand Project (Huhuanke2013-04). Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.chemosphere. 2015.07.051. References Arshadi, M., Soleymanzadeh, M., Salvacion, J., SalimiVahid, F., 2014. Nanoscale ZeroValent Iron (NZVI) supported on sineguelas waste for Pb (II) removal from aqueous solution: kinetics, thermodynamic and mechanism. J. Colloid Interface Sci. 426, 241–251. Badruddoza, A.Z.M., Shawon, Z.B.Z., Rahman, M.T., Hao, K.W., Hidajat, K., Uddin, M.S., 2013. Ionically modified magnetic nanomaterials for arsenic and chromium removal from water. Chem. Eng. J. 225, 607–615. Bhattacharya, A.K., Naiya, T.K., Mandal, S.N., Das, S.K., 2008. Adsorption, kinetics and equilibrium studies on removal of Cr(Vi) from aqueous solutions using different low-cost adsorbents. Chem. Eng. J. 137, 529–541. Bhattacharyya, K.G., Gupta, S.S., 2007. Adsorptive accumulation of Cd(II), Co(II), Cu(II), Pb(II), and Ni(II) from water on montmorillonite: influence of acid activation. J. Colloid Interface Sci. 310, 411–424. Chen, Z.X., Jin, X.Y., Chen, Z., Megharaj, M., Naidu, R., 2011. Removal of methyl orange from aqueous solution using bentonite-supported nanoscale zero-valent iron. J. Colloid Interface Sci. 363, 601–607. Cundy, A.B., Hopkinson, L., Whitby, R.L., 2008. Use of iron-based technologies in contaminated land and groundwater remediation: a review. Sci. Total Environ. 400, 42–51. Du, C.W., Li, X.G., Chen, X., Liang, P., Guo, H., 2008. Crevice corrosion behavior of X70 steel in HCO 61 3 solution under cathodic polarization. Acta Metall. Sin. 21, 235–244. Fei, X., Cao, L., Zhou, L., Gu, Y., Wang, X., 2012. Degradation of bromamine acid by nanoscale zero-valent iron (nZVI) supported on sepiolite. Water Sci. Technol. 66, 2539–2545. Geng, B., Jin, Z., Li, T., Qi, X., 2009. Kinetics of hexavalent chromium removal from water by chitosan-Fe0 nanoparticles. Chemosphere 75, 825–830.

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