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Mechanisms of Synergistic Removal of Low Concentration As(V) by nZVI@Mg(OH) Nanocomposite 2

Can Wu, Weizhen Liu, Jing Zhang, Shengqi Chu, Zhenqing Shi, Zhang Lin, and Zhi Dang J. Phys. Chem. C, Just Accepted Manuscript • DOI: 10.1021/acs.jpcc.7b06356 • Publication Date (Web): 06 Sep 2017 Downloaded from http://pubs.acs.org on September 8, 2017

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The Journal of Physical Chemistry

Mechanisms of Synergistic Removal of Low Concentration As(V) by nZVI@Mg(OH)2 Nanocomposite Can Wu,a,c Weizhen Liu,a,c Jing Zhang,b Shengqi Chu,b Zhenqing Shi,a Zhi Dang,a and Zhang Lin*a a School of Environment and Energy, South China University of Technology, The Key Laboratory of Pollution Control and Ecosystem Restoration in Industry Clusters (Ministry of Education), Guangdong Engineering and Technology Research Center for Environmental Nanomaterials, Guangzhou Higher Education Mega Center, Guangzhou, Guangdong 510006, China b Beijing Synchrotron Radiation Facility, Institute of High Energy Physics, Chinese Academy of Sciences, Beijing 100049, China c Equal contribution

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ABSTRACT: In this work, by using Mg(OH)2 nanoplatelets as support material for nanoscale zero valent iron (nZVI), nZVI@Mg(OH)2 composite was prepared and found with super high adsorption ability to As(V) at environmentally relevant concentrations. It revealed that the variation of corrosion products of nZVI in the presence of Mg(OH)2 and Mg2+ is an important factor for increase in adsorption ability to As(V). X-ray diffraction (XRD) analysis indicated that the weakly basic condition induced by Mg(OH)2 decreases the lepidocrocite (γ-FeOOH) and increases the magnetite/maghemite (Fe3O4/γ-Fe2O3) content in the corrosion products of nZVI, and the latter has better adsorption affinity to As(V). Moreover, extended X-ray absorption fine structure spectroscopy (EXAFS) indicated that the coordination between arsenic and iron minerals is influenced by dissolved Mg2+, leading to probable formation of magnesium ferrite (MgFe2O4) which has considerable adsorption affinity to As(V). This work provides an important reference not only for the design of pollution control materials but also for understanding arsenic immobilization in natural environments with ubiquitous Mg2+ ion.

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INTRODUCTION Arsenic (As) is a naturally-occurring element and a toxic metalloid that can adversely affect human health1. Natural As pollution in drinking water has been reported in more than 70 countries2. It is worth mentioning that the contamination of As in groundwater is generally at a low concentration, ranging from 0.01 to 5 mg/L in most countries3. The accepted concentration of As in drinking water has been reduced to 10 µg/L by WHO since 19934. Among numerous methods for As removal5, adsorption is most commonly adopted because it is inexpensive, generates little by-product, and does not require complex operation6. Nanoscale zero valent iron (nZVI) has been considered as a promising adsorbent because of its high reactivity, superior adsorption capacity and strong affinity to As7,8. The nZVI particles on their own tend to aggregate spontaneously, and to prevent such aggregation, they are generally supported on various materials to give a composite that possesses higher specific surface area of nZVI and thus better ability for As removal9,10,11,12,13. During the removal of As by nZVI, the As is mainly removed by adsorption to the iron corrosion products14. Iron corrosion products including lepidocrocite (γ-FeOOH), maghemite (γ-Fe2O3), and/or magnetite (Fe3O4), is with different adsorption ability towards As7,8. It has been reported that under same condition, magnetite shows higher adsorption ability to As than lepidocrocite15,16. The formation of the corrosion products of nZVI can be influenced by external conditions, such as the solution pH and the presence of Fe2+ or Fe3+ ion17 and the modifying agent18. Hence, in evaluating 3



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the ability of nZVI composites to remove As, besides considering the increase in specific surface area, it is necessary to determine whether the iron corrosion products change in the presence of supporting materials, and if so, how does the change affect the As adsorption ability of the composite. However, to the best of our knowledge, the influence of supporting materials on the corrosion products of nZVI has never been revealed. As an environmentally friendly material for water treatment, Mg(OH)2 shows very high adsorption affinity for both As(III) and As(V) due to the formation of Mg– As compound or Mg(OH)2–As(V) inner sphere complexes19,20,21,22. In our previous study, we prepared a novel flower-like nano-Mg(OH)2-supported nZVI that effectively suppressed the aggregation of nZVI particles23. We expect this novel nZVI@Mg(OH)2 composite to effectively adsorb As because of the following synergistic effects. Firstly, because Mg(OH)2 is weakly basic, it may change the corrosion products of nZVI, generation iron species with relatively strong affinity to As, and consequently enhance the As adsorption ability of the composite. Secondly, the dissolved Mg2+ ions from Mg(OH)2 may incorporate into the structure of magnetite and displace Fe2+ to generate magnesium ferrite, which has been recently reported to have stronger adsorption ability to As than many other iron oxides, including magnetite24. Up to now, the influence of Mg2+ (which is a ubiquitous ion in environment) on the interaction between iron minerals and arsenic has never been revealed. In this work, we investigated As(V) removal efficiency and related mechanism of 4


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nZVI@Mg(OH)2 in comparison with nZVI and Mg(OH)2 at environmentally relevant concentrations, and further examined the synergistic mechanism between nZVI and Mg(OH)2 in the removal of As(V). As(V) removal by A series of spectroscopic and microscopic methods, including X-ray diffraction (XRD), transmission electron microscopy (TEM), X-ray photoelectron spectroscopy (XPS), and X-ray absorption fine structure (XAFS), were conducted. The objectives of this study are to (1) quantify the As(V) adsorption ability of nZVI@Mg(OH)2 at low concentrations, (2) investigate the transformations of the iron phases during the adsorption process, and (3) assess the role of Mg(OH)2 in the corrosion of nZVI and its impact on the removal of As(V). This study will provide an important reference not only for the design of pollution control materials meet the actual demand in environment, but also for understanding the stability of arsenic in natural environments.

MATERIALS AND EXPERIMENT Chemicals. All the chemicals and reagents in this study were of analytical grade. The stock solutions of As(V) were prepared by dissolving Na2HAsO4·7H2O in deionized water. Synthesis of Adsorbents. Adsorbents were prepared according to the method described in our previous study23. Specifically, Mg(OH)2 was prepared at room temperature by precipitation from NaOH and MgSO4·7H2O solutions, nZVI was prepared by reducing FeSO4 with NaBH4 in aqueous solution, and nZVI@Mg(OH)2 was prepared by the in-situ synthesis of nZVI in a suspension of Mg(OH)2. Detailed 5



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procedures can be found in the supplementary information (SI). Batch Adsorption Experiments. As(V) adsorption experiments were conducted with nZVI, Mg(OH)2 and nZVI@Mg(OH)2. All experiments were run at room temperature and in open air, and the concentration of adsorbent was fixed at 1g/L. To study the equilibrium time for As(V) adsorption, kinetic experiments were run up to 5 hours and sampled at selected time intervals, with an initial pH of 7.0 ± 0.1 and a total As(V) concentration of 5 mg/L. For the rest of adsorption experiments, the reaction time was fixed at 5 hours. To investigate the As(V) adsorption capacity, the total As(V) concentrations (0.5 – 5 mg/L) were varied in selected batch adsorption experiments, with an initial pH of 7.0. In addition, to understand how the pH affected the corrosion products of nZVI during the As(V) adsorption, nZVI was reacted with 5 mg/L As(V) solution with different initial pH (4.0, 7.0 and 10), and the corrosion products were collected and analyzed. For all experiments, no attempt was made to maintain a constant pH once the adsorbents were added and the variation of the pH value during As(V) adsorption was recorded. The sampled suspensions were filtered using a 0.22 µm filter and the concentrations of As, Fe and Mg in solutions were measured by inductively coupled plasma mass spectrometry (ICP-MS). All experimental details can be found in the SI. Characterization. XRD data were collected on a Bruker X-ray powder diffractometer (advance D8) with Cu-K α-radiation. The tube voltage was 40 kV and the tube current was 40 mA. The specific surface area of the samples was measured by the Brunauer– Emmett–Teller (BET) N2 adsorption method on a Micromeritics 6


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Gemini V2.00 (Model 2380) instrument, in which the samples were dried with a constant flow of N2 at 100 °C for 6 h. Images of adsorbents were taken under a JSM-7100F scanning electron microscope (SEM) with an Oxford-INCA energy dispersive X-ray spectroscopy (EDS) instrument and a TEM (JEOL JEM2010). The element composition and valence state of solid phases were determined by XPS (ESCAL AB250) with Al-Kα radiation (1487 eV) at 300 W under of 1.0 × 10−5 Pa. To study the local molecular environment of As-loaded adsorbents, the XAFS spectra of As K-edge and Fe K-edge were acquired at the beamline 4W1B of the Beijing Synchrotron Radiation Facility (BSRF), China. The details of XAFS spectra collection and data analysis are presented in the SI.

RESULTS AND DISCUSSION Characterizations of Adsorbents. Similar to what we previously reported23, the XRD pattern of the nZVI@Mg(OH)2 sample contains the diffraction information of both nZVI and Mg(OH)2 phases (SI Figure S1). ICP-MS analysis shows that the Fe content in nZVI@Mg(OH)2 is 33.66%. According to the BET measurements, the specific surface area of nZVI, Mg(OH)2 and nZVI@Mg(OH)2 is 21.6, 36.69, and 51.22 m2/g, respectively. The data indicate good dispersion of nZVI on Mg(OH)2. TEM and SEM observation reveal that the nZVI nanoparticles tend to aggregation spontaneously25 (Figure 1a). Mg(OH)2 has a self-supported spherical structure, whose diameter is of several micrometers (SI Figure S2), interweaved by nanoplatelets of 8– 20 nm thickness (Figure 1b)26. The nZVI@Mg(OH)2 contains 30–70 nm nanoparticles 7



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homogeneously distributed on the surface of nanoplatelets (Figure 1c). Further EDS elemental mapping of nZVI@Mg(OH)2 (Figure 1d) shows that the Mg-K and O-K signals appear homogeneously on all nanoplatelets, whereas the position of the Fe-K signals match well with the nanoparticles, indicating that nZVI nanoparticles are well distributed on the surface of Mg(OH)2. Therefore, the results from XRD, BET, SEM and TEM analysis, collectively, confirm that the spontaneous aggregation of nZVI nanoparticles is effectively prevented when they are supported on the flower-like Mg(OH)2.

Figure 1. (a) TEM image of nZVI, scale bar = 200 nm. (b) SEM image of flower-like nano-Mg(OH)2, scale bar = 300 nm. (c) SEM image of nZVI@Mg(OH)2 composite, scale bar = 200 nm. (d) High angle annular dark field scanning transmission electron 8


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microscopy (HAADF–STEM) image of nZVI@Mg(OH)2 and the corresponding EDS elemental mapping images, scale bar = 200 nm.

Batch Experiments of As(V) Adsorption. Figure 2 shows that as the initial concentration C0 of As(V) decreases from 5 to 0.5 mg/L, the residual concentration Ce of As(V) decreases accordingly for all three samples. Over the entire As(V) concentration range, the nZVI@Mg(OH)2 nanocomposite always shows the strongest As(V) removal ability, which indicates its superior adsorption affinity to As(V). Noted that when the initial concentration C0 of As(V) was 1 mg/L, the residual concentration Ce of As(V) in the solutions treated by nZVI, Mg(OH)2 and nZVI@Mg(OH)2 is 18.7, 7.9 and 3.6 µg/L respectively. It reveals a superior ability of nZVI@Mg(OH)2 to remove As(V) from water to meet the WHO standard (10 µg/L).

Figure 2. The residual concentration Ce of As(V) after the treatment with nZVI, Mg(OH)2 and nZVI@Mg(OH)2 for 5 h at varying initial concentrations (C0). Initial concentration of As(V) was 0.5 − 5 mg/L; dose of adsorbents was 1 g/L; the solution volume was 20 ml; and pH was 7.0 ± 0.1. 9



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The equilibrium adsorption data and related isotherm parameters of nZVI, Mg(OH)2 and nZVI@Mg(OH)2 to As(V) were analyzed using both Langmuir and Freundlich isotherm models (Figure 3). As show in SI Table S1, nZVI sample has the higher correlation coefficient (R2) value for Langmuir adsorption model. The adsorption isotherm of As(V) by Mg(OH)2 sample fitted well with both Freundlich and Langmuir adsorption models. Regarding that Langmuir adsorption model is normally used to describe the monolayer adsorption and Freundlich adsorption model is use to describe the multilayer adsorption, due to the As in solution is at extremely low concentration, we majorly considered the fitting results from Langmuir adsorption mode. Fitting results for nZVI and nZVI@Mg(OH)2 also indicated Langmuir model was more appropriate to depict the adsorption process of As(V) on these materials, which was consistent with previous reports of adsorption As(V) on nZVI7,27 and Mg(OH)219. The parameters Kl calculated from Langmuir adsorption isotherm of nZVI, Mg(OH)2 and nZVI@Mg(OH)2 to As(V) was 3.15, 3.33 and 17.24 L/mg, respectively. As we know, Kl is a constant related to the energy of adsorption, which represents the affinity between adsorbate and adsorbent, and reflects the binding strength28. The above result implied that the As(V) adsorption affinity was remarkably enhanced by using Mg(OH)2 as supporting material for nZVI.

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Figure 3. Adsorption isotherms for As(V) on (a) nZVI, (b) Mg(OH)2 and (c) nZVI@Mg(OH)2. Initial concentration of As(V) was 1 − 8 mg/L; dose of adsorbents was 1 g/L; the solution volume was 20 ml; and pH was 7.0 ± 0.1.

Adsorption kinetic is a key factor for investigate the adsorption rate of adsorbent to contamination. The As(V) uptake by nZVI@Mg(OH)2 gradually reached equilibrium within about 60 min (Figure 4a) and the time to equilibrium is relatively faster than both nZVI and Mg(OH)2. The adsorption kinetic data of nZVI, Mg(OH)2 and nZVI@Mg(OH)2 to As(V) were analyzed by using both pseudo first order and pseudo second order models (Figure 4b and c) and the related parameters are listed in SI table S2. The higher correlation coefficient (R2) value of pseudo second model for the above three materials indicated that this model is more appropriate to depict the adsorption process of As(V) on these materials. 11



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Figure 4. Adsorption kinetics for As(V) on nZVI, Mg(OH)2 and nZVI@Mg(OH)2, (a) the adsorption capacity of As(V) at interval time, (b) fitted with pseudo first order model and (c) fitted with pseudo second order model. Initial concentration of As(V) was 5 mg/L; dose of adsorbents was 1 g/L; the solution volume was 20 mL; and pH was 7.0 ± 0.1. Mechanism of As(V) Removal by nZVI@Mg(OH)2. (1) The Changes of nZVI@Mg(OH)2 during Adsorption. XPS analysis of the elemental valence state of nZVI@Mg(OH)2 before and after As(V) adsorption (SI Figure S3 and S4) shows that As(V) is adsorbed on nZVI@Mg(OH)2 without reduction but Fe0 is oxidized during the adsorption. It indicated that water and oxygen are the primary oxidants for nZVI but As(V) cannot directly oxidize nZVI29. Although it was demonstrated that As(V) can be reduced to As(III) and As(0) by nZVI under anaerobic condition30, the reduction of As(V) cannot take place under aerobic 12


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conditions in this work, which is consistent with previous study27,29. Arsenic K-edge XANES spectra (SI Figure S5) of As-loaded nZVI, Mg(OH)2 and nZVI@Mg(OH)2 further confirmed that there is no reduction of As(V) in this study. The ICP-MS results of the batch experiments (SI Figure S6) further indicate that Mg is quickly dissolved into solution and reaches an equilibrium of 5 mg/L, whereas Fe does not readily dissolve and As(V) is effectively removed from the solution. The pH value of the solution increases rapidly at the beginning of As(V) adsorption and finally reaches a plateau (about 9) when the adsorption arrives at equilibrium (SI Figure S7). Since the eventual solution pH is only increase to 7.7 when nZVI was used for removal As(V) at initial pH = 7, the higher pH here must have result from the dissolution of Mg(OH)2. (2) Phase Transformation of the Fe Component during Adsorption. The weak basic solution that resulted from dissolution of Mg(OH)2 may influence the corrosion products of the nZVI component in the nanocomposite. To verify this, we first ran adsorption experiments with nZVI under different initial solution pH and analyzed the corrosion products. At initial pH = 4, upon interaction with As(V) for 5 h (Figure 5b), nZVI gives corrosion products consisting of maghemite/magnetite (γ-Fe2O3/Fe3O4), lepidocrocite (γ-FeOOH) and goethite (α-FeOOH). In contrast, at initial pH = 7 (Figure 5c) the corrosion products consist of maghemite/magnetite (γ-Fe2O3/ Fe3O4) and lepidocrocite (γ-FeOOH)7, and at pH = 10 (Figure 5d) the corrosion products become predominantly maghemite/magnetite (γ-Fe2O3/Fe3O4). Clearly, the corrosion products of nZVI vary with the initial solution pH. Specifically, acid solution tends to 13



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give

lepidocrocite

(γ-FeOOH),

whereas

basic

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solution

mainly

gives

maghemite/magnetite (γ-Fe2O3/Fe3O4). We then analyzed the corrosion products of nZVI@Mg(OH)2 after As(V) adsorption for 5 h at initial pH of 7. The main phases of samples

after

reaction

include

Fe0,

Mg(OH)2

and

maghemite/magnetite

(γ-Fe2O3/Fe3O4) (Figure 5f), which coincide with the corrosion products of nZVI at pH = 10. It indicated that the coexisting Mg(OH)2 has altered the component of nZVI corrosion products in the composite. Since magnetite/maghemite has stronger adsorption affinity to As(V) than the lepidocrocite, the As(V) removal ability of the composite may improves accordingly16.

Figure 5. XRD patterns of (a) nZVI before As(V) adsorption, (b) As(V)-loaded nZVI 14


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at initial pH = 4, (c) As(V)-loaded nZVI at initial pH = 7, (d) As(V)-loaded nZVI at initial pH = 10, (e) nZVI@Mg(OH)2 before As(V) adsorption, and (f) As(V)-loaded nZVI@Mg(OH)2. (●, Fe0; ■, Mg(OH)2; ▽, maghemite/magnetite, γ-Fe2O3/Fe3O4; ◇, lepidocrocite, γ-FeOOH; ★, goethite, α-FeOOH; adsorbent dose, 1 g/L; adsorption time, 5h.)

It is also worth noting that Mg(OH)2 is slightly soluble in water, it is possible to generate magnesium ferrite (MgFe2O4) in the presence of Mg2+. Because the XRD spectrum of magnesium ferrite (MgFe2O4) has the same peaks as that of maghemite/magnetite (γ-Fe2O3/Fe3O4), it is difficult to exclude the existence of magnesium ferrite (MgFe2O4) in As(V)-loaded nZVI@Mg(OH)2. Therefore, it is prudent to conclude that under the influence of Mg(OH)2, the corrosion product of nZVI consists of maghemite(γ-Fe2O3), magnetite (Fe3O4), and probably magnesium ferrite (MgFe2O4). Once magnesium ferrite existed, based on the fact that it has stronger adsorption affinity to As(V) than magnetite/maghemite24, the adsorption ability of the composite could be improved further. (3) EXAFS Analysis of the Coordination of As(V). The Above results show that (1) the weak basic condition caused by Mg(OH)2 affects the corrosion products of nZVI and (2) the Mg2+ in the solution may led to the formation of magnesium ferrite. The following EXAFS analysis of the coordination of As(V) further elaborates on how these two factors enhance the adsorption ability of nZVI@Mg(OH)2. To assess the pH contribution caused by Mg(OH)2, the arsenic K-edge in the EXAFS data of As(V)-loaded nZVI@Mg(OH)2 is compared with that of 15



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As(V)-loaded nZVI prepared at different initial solution pH (i.e., 4, 7, and 10). Figure 6 shows the unfiltered k3-weighted As K-edge EXAFS data and their Fourier transforms, and SI Table S3 lists the corresponding fitting parameters of the unfiltered k3x(k) EXAFS functions. It can be seen from Figure 6 that for all samples, the Fourier transformed R-space spectra consist of a predominant signal of As–O first-neighbor contributions and a weaker signal of As–Fe and/or As–Mg second-neighbor contributions. Since for all samples the As–O first-neighbor contributions can be fitted with 3.9–4.1 oxygen atoms at 1.69 ± 0.02 Å, which correspond to the molecular structure of AsO4 tetrahedron, the following analysis will focus on the analysis of second-neighbor contributions. In As(V)-loaded nZVI@Mg(OH)2, the signal of second-neighbor contributions is relatively weaker than that of pure iron oxides samples, which can be due to the presence of magnesium. Such trend is also confirmed by aluminum doped31. To estimate which atoms (Fe or Mg) will most likely backscatter in the second shell, the EXAFS spectra are fitted together using the As–Fe of FeAsO4·2H2O and the As–Mg paths of NaMg4(AsO4)3. The second-neighbor contributions can be fitted by 2.0 As– Fe pairs and 1.9 As–Mg pairs at 3.41 Å and 3.38 Å, respectively. Therefore, As can simultaneously combine with Fe and Mg through inner-sphere complexation. The As– Fe pair at 3.41 Å comes from the bidentate binuclear double-corner (2C) surface complexes, and the observed distance is close to the As–Fe distance assigned to the 2C complexes for As(V) adsorption onto maghemite (3.38 ± 0.06 Å)32. The As–Mg pair at 3.38 Å can be attributed to the formation of monodentate mononuclear (1V) 16


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surface complexes, and the result is consistent with the As–Mg pair assigned to the 1V complexes for As(V) adsorption onto Mg(OH)2 (3.38 ± 0.02 Å)19.

Figure 6. XAFS spectra of As(V)-loaded nZVI obtained at different initial solution pH and As(V)-loaded nZVI@Mg(OH)2. (a) Unfiltered k3-weighted k-space, (b) magnitude part of the Fourier transformed R-space, and (c) real part of the Fourier transformed R-space. Red open circles and black solid lines represent experiment data and fitted curve, respectively.

As we know, the pH may influence the adsorption affinity of the nZVI@Mg(OH)2 to As(V) since increasing OH– will has electrostatic repulsive-force to As anion. For evaluation the contribution of pH, we tried to get the Fe/As coordination number for As(V)-loaded nZVI at different initial pH (i.e. 4, 7, and 10) as comparing results. For the nZVI samples obtained after As(V) adsorption at different initial solution pH, the second-neighbor contributions to the EXAFS are fitted by using As–Fe pairs at various distances. For all three samples, the As atom is 17



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surrounded by Fe atoms at 3.33–3.36 Å, which suggests the formation of bidentate binuclear corner-sharing (2C) complexes. The As-Fe distance is close to previously reported values for As(V) adsorption onto magnetite33, maghemite34 and lepidocrocite35,36. However, the coordination number of Fe atoms around the As atom decreases with rising initial pH value. Specifically, the Fe/As coordination number is 1.5, 1 and 0.8 for As(V)-loaded nZVI samples prepared at an initial solution pH of 4, 7, and 10 respectively (in which the final pH is 6.7, 7.7 and 8.9). This is because when the initial pH increases from 4 to 10, even though the magnetite/maghemite phase (with strong affinity to As) increases, the positive charge on the adsorbent surface decreases due to increasing OH– content, and as a result the As(V) sorption is reduced37. In contrast, the As(V)-loaded nZVI@Mg(OH)2 has a high Fe/As coordination number of 2, much higher than that of As(V)-loaded nZVI prepared at initial pH = 10. That is to say, under the same weakly basic condition, the nZVI@Mg(OH)2 composite has much higher adsorption affinity to As(V) than nZVI. Therefore, besides pH value, some other factor must have also contributed to the strong adsorption ability nZVI@Mg(OH)2 for As(V). In the following discussion, we examine whether magnesium ferrite has improved the adsorption ability of nZVI@Mg(OH)2 as well.

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Figure 7. Fe K-edge data recorded for nZVI@Mg(OH)2 and for nZVI at pH = 10. (a) (c) Unfiltered k3-weighted k-space. (b) (d) Magnitude and real part of the Fourier transformed R-space. Red open circles and black solid lines represent experimental data and fitted curve, respectively.

First, the Fe K-edge EXAFS data of nZVI@Mg(OH)2 are compared with that of nZVI prepared at pH = 10. The raw data are Fourier filtered to isolate the peaks between 1 and 3 Å in the real space and back-transformed to yield the EXAFS pattern shown in Figure 7. The fitted data are given in Table S4. For the nZVI, the Fe K-edge EXAFS data are fitted by using the structure model of magnetite (Figure 7c, d). The fitting indicates 6 O atoms around Fe at 1.97 Å and 4.6 Fe atoms around Fe at 2.99 Å, respectively. The results conforms to the typical octahedral structure of FeO6 and the structure of edge-sharing Fe–Fe or corner-sharing Fe–Fe38. In contrast, the EXAFS data and fitting result of nZVI@Mg(OH)2 are obviously different. It has been 19



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previously reported that magnesium ferrite is a mixed Fe/Mg oxide with an inverse spinel structure, in which the Fe ions occupy both tetrahedral (Td, Fe-O =1.92 Å) and octahedral (Oh, Fe–O = 2.04 Å) sites38. Therefore, the data of nZVI@Mg(OH)2 are fitted into a simple three shell model of magnesium ferrite, and the final parameters are listed in SI Table S4. The second-neighbor contributions can be fitted by 1.9 Fe– Mg pairs at 3.26 Å, which indicates that Mg2+ can be incorporated into the structure of iron oxide and form of magnesium ferrite. This result is consistent with the previous study39.

Figure 8. Proposed surface complex models of As(V) adsorbed on nZVI@Mg(OH)2. From the EXAFS results and the crystal structure model of magnesium ferrite40 (SI Figure S8), we propose three types of surface complexes of As(V) on nZVI@Mg(OH)2: (1) As–Fe bidentate binuclear surface complexes (As–Fe distance = 3.41 Å); (2) As–Mg monodentate mononuclear surface complexes (As–Mg distance = 3.38 Å); (3) possible bidentate binuclear surface complexes between As, Fe and Mg. Figure 8 illustrates these structures.

CONCLUSIONS 20


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To sum up, the synergy between the following factors contribute to the superior performance of the nZVI@Mg(OH)2 in adsorbing As(V). First, Mg(OH)2 can not only acts as an excellent supporting material to prevent the aggregation of nZVI but also provides strong adsorption affinity to As(V). Second, the hydroxides provided by Mg(OH)2 form a weakly basic environment around nZVI. Consequently, in the corrosion

products

of

nZVI,

less

lepidocrocite

is

generated

but

more

magnetite/maghemite is produced. Since magnetite/maghemite has stronger adsorption affinity to As(V) than the lepidocrocite, the performance of the composite improves accordingly16. Last, dissolution of Mg(OH)2 provides Mg2+ in the solution, which may be incorporated into the structure of magnetite, displace the position of Fe2+, and generate of magnesium ferrite (MgFe2O4). Since magnesium ferrite has stronger adsorption affinity to As(V) than magnetite/maghemite23, the adsorption ability of the composite is further improved. As is normally present in polluted groundwater at a low concentration (0.01 – 5 mg/L), it is desirable to employ adsorbents with excellent adsorption affinity rather than those with excellent adsorption capacity. On the other hand, the interaction between arsenic and iron oxides is a very important issue relating to the pollution control of arsenic in water as well as the immobilization of arsenic in natural environment. In this work, efforts were committed to disclose the excellent adsorption affinity of nZVI@Mg(OH)2 for removal As(V). To our knowledge, this is the first study to assess the important role of Mg2+ ion (a main element in the earth) on the iron oxides for immobilizing As(V). Also, we hope our results shed new light on the 21



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understanding of the important role of supporting materials, in which the corrosion products of nZVI could be adjusted for achieving a higher adsorption affinity to remove low concentration of contaminations in the environment.

ASSOCIATED CONTENT Supporting Information Additional data and experimental details are presented in Supporting Information section.

AUTHOR INFORMATION Corresponding Author * E-mail: [email protected]; Tel: 86-20-39380503; Fax: 86-20-39380508. Notes The authors declare no competing financial interest. ACKNOWLEDGMENTS This work was supported by the National Natural Science Foundation of China (grant no.21477129 and 21273237), the Outstanding Youth Fund (no. 21125730), and the Fundamental Research Funds for the Central Universities (no. 2015ZM157). The authors also thank the beamline 4W1B (Beijing synchrotron radiation facility) for providing the beam time.

22


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