Simultaneous photooxidation and sorptive removal of As(III) by TiO2 supported layered double hydroxide

Journal of Environmental Management 161 (2015) 228e236

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Research article

Simultaneous photooxidation and sorptive removal of As(III) by TiO2 supported layered double hydroxide Sang-Ho Lee a, Kyoung-Woong Kim a, *, Heechul Choi a, Yoshio Takahashi b a b

School of Environmental Science and Engineering, Gwangju Institute of Science and Technology, Gwangju, Republic of Korea Department of Earth and Planetary Science, Graduate School of Science, The University of Tokyo, Tokyo, Japan

a r t i c l e i n f o

a b s t r a c t

Article history: Received 23 April 2015 Received in revised form 24 June 2015 Accepted 26 June 2015 Available online 15 July 2015

The present study focused on the enhanced removal of As(III) by the simultaneous photooxidation and removal process using TiO2 nanoparticles supported layered double hydroxide (TiO2/LDH). The TiO2/LDH nanocomposites were synthesized using a flocculation method, and nanosized (30e50 nm) TiO2 particles were well-distributed on the LDH surface. The XPS and DLS data revealed that the TiO2/LDH nanocomposites were both chemically and physically stable in the aquatic system. The optimum ratio of TiO2 was 20 wt.% and the calcination process of LDH enhanced the removal capacity of As(III) by the reconstruction process. In the kinetic removal experiment, UV irradiation improved the removal rate of As(III), based on the continuous conversion of As(III) to As(V), and that the removal rate was faster under alkaline conditions than acidic and neutral conditions due to the abundance of oxidants and negative charged As(III) species (pKa: 9.2). The main mechanism of As(III) photooxidation is the direct oxidation by hvb þ, which is generated by supported TiO2 nanoparticles. X-ray near edge structure results also confirmed that the As(III) was completely oxidized to As(V). Consequently, the simultaneous photooxidation and removal process of As(III) by TiO2/LDH nanocomposites may be the effective removal option in As(III) contaminated water. © 2015 Elsevier Ltd. All rights reserved.

Keywords: Arsenite TiO2/LDH nanocomposites Photooxidation Reconstruction process X-ray absorption near edge structure

1. Introduction Arsenic (As) is a toxic element that has also been classified as a strong carcinogen (Bissen and Frimmel, 2003). In general, arsenic exists in four oxidation states i.e., þ5, þ3, 0, and 3, with trivalent arsenic [arsenite; As(III)] and pentavalent arsenic [arsenate; As(V)] being the dominant arsenic species in an aquatic environment (Smedley and Kinniburgh, 2002). Arsenic in the environment can be originated from both natural and anthropogenic sources; nevertheless, naturally occurring sources negatively impact the aquatic ecosystem such as in the contamination of ground and surface water (Chapagain et al., 2009; de Figueiredo et al., 2007; Yang et al., 2009). Recently, severe arsenic contamination in drinking water has been observed around the world, making the chronic exposure of arsenic a global issue due to increases in kidney, lung, bladder, and skin cancer (Anawar et al., 2001; Kurttio et al., 1999; Smith et al., 2000, 1992). In fact, previous studies have reported that the people who drink the arsenic contaminated

* Corresponding author. E-mail address: [email protected] (K.-W. Kim). http://dx.doi.org/10.1016/j.jenvman.2015.06.049 0301-4797/© 2015 Elsevier Ltd. All rights reserved.

water display considerable arsenic accumulation in their body, such as in their hair, nails, toenails, and urine showing the positive correlation with the contamination level of their groundwater (Phan et al., 2010, 2013). From the risk of arsenic contamination in drinking water, the US Environmental Protection Agency (USEPA), World Health Organization (WHO), and European Commission have set to the maximum level of arsenic in drinking water to be 10 mg/L (10 ppb) (Environmental Protection Agency, 2001; WHO, 2011). Representative arsenic removal techniques in drinking water include the precipitation/coagulation, adsorption, ion exchange processes and membrane treatments such as reverse osmosis and nanopore filtration systems. In particular, adsorption and ion exchange processes have been widely applied to the contaminated areas due to their cost-effectiveness and installation convenience (Mohan and Pittman, 2007). A common problem in arsenic removal systems is that the removal efficiency of As(III) is lower than that of As(V), that is because As(III) exists in a non-ionic form as H3AsO3 in neutral conditions, and therefore has a higher mobility than As(V) species, which exist in negative charged forms such as H2 AsO4  , HAsO4 2 , and AsO4 3 (Lin and Wu, 2001). For this reason, oxidation

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processes have been applied as a pretreatment in arsenic removal systems in attempts to enhance the overall removal efficiency. To oxidize As(III) in the pretreatment system, advanced oxidation processes such as O3 and H2O2 and TiO2-mediated photooxidation process have been applied (Neppolian et al., 2009; Pettine et al., 1999; Ryu and Choi, 2006). The application of nanosized TiO2 suspension in conjunction with UV irradiation has been a promising remediation technique to oxidize As(III) to As(V) in treatment systems. However, the direct application of nanosized TiO2 particles can be still challenging problem due to the difficulty of separation. To date, several researchers have focused on the immobilization of nanosized TiO2 particles on substrates such as adsorbents (Li et al., 2014), electrodes (Pifferi et al., 2013) and membranes (Zhang et al., 2015). In recent, TiO2 immobilization techniques have been applied to arsenic adsorbents such as iron oxide and aluminum oxide in order to synthesize bi-functional nanocomposites to induce the simultaneous photocatalytic oxidation and adsorption process (Yu et al., 2013; Zhou et al., 2008). The layered double hydroxide (LDH) is the anionic clay that is based on the staking of positively charged brucite-like layers with interlayers of anion and water molecules. The LDHs can be denoted by the general formula ½MII 1x MIII x ðOHÞ2 xþ ðAn Þx=n $mH2 O, where MII and MIII are divalent and trivalent cations, respectively, x is equal to the molar ratio of MIII(MII þ MIII), A is an interlayer anion having charge m, and n is the number of water molecules (Goh et al., 2008). The LDH has the large surface area and high anion exchange capacity, and has been applied to remove the anionic contaminants such as arsenic (Türk et al., 2009), phosphate (Cheng et al., 2009), chromium (Das et al., 2004) and antimony (Kameda et al., 2012) based on the mechanisms of adsorption, anion exchange in the interlayer region, and the reconstruction of a calcined LDH structure. Furthermore, LDHs have been combined with materials such as graphene (Li et al., 2010), polymers (Cho et al., 2013), and inorganic metals (Jiao et al., 2009) to enhance their reactivities by maximizing their inherent properties (e.g., large surface area and layered structure). Previous research reported that LDHs have been applied as substrates to the synthesize TiO2/LDH nanocomposites for the  et al., 2012; Shao photodegradation of organic pollutants (Pau
sova et al., 2014), and TiO2/LDH nanocomposites increased the photodegradation of organic pollutants by activating hydroxyl groups on the LDH surface (Mendoza-Dami an et al., 2013). This study presents the results obtained by combining LDH and TiO2 to remove As(III) using a simultaneous photooxidation and removal process in drinking water. The TiO2/LDH nanocomposites are the effective bi-functional material that has both the higher oxidation and removal efficiency of arsenic in comparison with other commercial adsorbents. Therefore, this TiO2/LDH nanocomposites could be an effective nanocomposites for environmental applications in targeting for the As(III) contaminated areas. 2. Materials and methods Carbonate type LDH was synthesized by the modified method with previous study (Paredes et al., 2011). In brief, Soln. 1 (1 M NaHCO3 with 3.5 M NaOH) was slowly added into Soln. 2 (1 M Mg(NO3)2 6 $ H2O and Al(NO3)3 9H2O) using drop-wise method until the pH of the mixed solution reached 9.5. After mixing, the solution was aged for 24 h at 60  C to increase its crystallinity. The precipitate was then centrifuged at 6000 rpm and washed using deionized water. The samples were dried at 80  C in a dry oven for 24 h and finally stored in a desiccator. Anatase type titanium dioxide was synthesized using the hydrolysis method described by Yun et al. (2012). In brief, 50 mL of

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0.5 M TiCl4 (99.9%; Sigma Aldrich, USA) was prepared and the titanium solution was mixed with ammonium sulfate ((NH4)2SO4; Sigma Aldrich, USA) to prepare the anatase TiO2. The mixed solution was constantly stirred at 110  C for 24 h in the reflux system. Subsequently, the produced precipitate was separated from the solution and washed with deionized water to remove excess ions such as chloride and sulfate. Finally, the precipitate was dried at 80  C and stored in a desiccator. The process for synthesizing TiO2/LDH nanocomposites was previously described by Seftel et al. (2010). Here, 0.8 g of LDH and 0.2 g of well-dispersed TiO2 nanoparticles were mixed into a 1 L bottle containing deionized water. The mixed solution was vigorously stirred for 48 h at room temperature. Prior to stirring the mixed solution, N2 gas was purged to prevent carbonate contamination. The final slurry was centrifuged and calcined at 400  C in a furnace, and finally filtered using a 100 mesh sieve. For the characterization, powder XRD patterns were measured using a high resolution X-ray diffractometer (HR-XRD) with Cu Ka radiation (40 KV, 24 mA) over a 2q range of 5 e85 . The BET surface area was analyzed using an ASAP2010 (Micromeritics, USA), and the morphological characteristics and elemental composition of the samples were obtained using scanning electron microscope (FESEM; Hitachi S-4700, Japan). The TiO2/LDH surface was analyzed using the X-ray photoelectron spectrometer (XPS; ESCA VG Multilab 2000, UK) to determine the differences in elemental characteristics on the surface of TiO2/LDH nanocomposites. The IR spectra were obtained using FT-IR (Jasco-4600 plus; USA), and arsenic Kedge X-ray absorption near-edge structure (XANES) spectra were obtained from the BL6A of a photon factory (High-Energy Accelerator Research Organization, KEK, Japan). The six UV lamps (NEC CFL3UV37, Japan; output: 5.2 mW/cm2; wavelength: 368 nm) were mounted in a shaking incubator (HB201SS, Hanbaek Science, Korea) equipped with a constant temperature circulator. The 30 mL of As(III) solution (10e50 mg/L) was mixed with TiO2/LDH nanocomposites and vigorously stirred at 180 rpm under UV irradiation. The solution was filtered using a 0.45 mm-sized membrane filter and the filtered solution was then passed through an anion exchange cartridge (Supelco LC-SAX SPE 57071, USA) to separate the As(III) species in the solution. The arsenic concentration in the stored solution was subsequently analyzed by inductively coupled plasma mass spectrometry (ICPMS; Agilent 7500 ce, USA). 3. Results and discussion 3.1. Characterization of TiO2/LDH nanocomposites Fig. 1 shows the XRD patterns for LDH, TiO2, and TiO2/LDH nanocomposites synthesized by spontaneous flocculation after mixing LDH with TiO2 nanoparticles. The XRD data revealed that uncalcined LDH shows typical structures of LDH (Fig. 1a), with basal reflection (003) and (006) at low 2q corresponding to the layered structure. Generally, the calculated thickness of the LDH interlayer is 7.65 Å (carbonate z 2.85 Å; layer thickness z 4.80 Å), which almost corresponds to the d-spacing value of the uncalcined LDH (7.80 Å) (Lu et al., 2012; Seftel et al., 2013). After the calcination process at 400  C, the typical LDH peaks disappeared due to the structural collapse caused by the loss of interlayer anions and water molecules incurred during the thermal treatment. Characteristics of the anatase phase (101) of TiO2 were generally observed at approximately 25 , with both uncalcined and calcined TiO2 displaying the typical anatase crystalline phase. The XRD pattern of uncalcined TiO2/LDH (Fig. 1c) includes both properties of LDH and TiO2 in basal planes (003), (006) and (101) planes. Interestingly, the calcined TiO2/LDH (Fig. 1d) shows that the (003) reflection

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Fig. 1. Powder X-ray diffraction pattern of uncalcined LDH (a), calcined LDH (b), uncalcined TiO2/LDH nanocomposites (20%; (c)), calcined TiO2/LDH nanocomposites (20%; (d)), uncalcined TiO2 (e), and calcined TiO2, (f) (* and ** indicate the peak corresponding to LDH and anatase TiO2, respectively).

disappeared after the thermal treatment, even though uncalcined TiO2/LDH has a small peak of (003) reflection. This finding indicates that excessive TiO2 nanoparticles possess the layer structure of the TiO2/LDH nanocomposites after the elimination of water molecules and interlayer anions. Therefore, the calcined TiO2/LDH lose the stacking periodicity of the layered structure; hence, the XRD patterns of calcined TiO2/LDH nanocomposites are similar to those of  et al., 2012). The previous research explained anatase TiO2 (Pau
sova that calcination process of LDH enhances the anion removal capacity by rehydration and sorption process (Goh et al., 2009), and the BET data also showed that the specific surface area of uncalcined TiO2/LDH nanocomposites (159.3 m2/g) slightly enhanced by calcination process (194.8 m2/g). Consequently the calcined TiO2/ LDH nanocomposites have the potential to remove anions by the reconstruction process. Fig. 2 displays the FT-IR spectra of LDH, TiO2, and TiO2/LDH nanocomposites, in which absorption bands at 3450, 1640 and 1380 cm1 were observed in the uncalcined LDH. The band center at 3450 cm1 indicates the characteristics of stretching vibrations

Fig. 2. FT-IR spectra of (a) uncalcined LDH, (b) calcined LDH, (c) uncalcined TiO2/LDH nanocomposites (20%), (d) calcined TiO2/LDH nanocomposites (20%), (e) uncalcined TiO2, (f) calcined TiO2.

in hydroxyl groups (-OH) in the brucite layers and interlayer water molecules, and the band at 1380 cm1 can be assigned to the asymmetric stretching of interlayer carbonate anions (CO3  ) (Lu et al., 2012). The calcined LDH demonstrated that hydroxyl group almost disappeared after the thermal treatment, as incomplete peaks of carbonate and nitrates were observed. Uncalcined TiO2/ LDH revealed that the hydroxyl and carbonate groups were restored due to the preparation method under flocculation in water, whereas these peaks were eliminated in calcined TiO2/LDH after the thermal treatment. It is indicated that there were no valid peaks in the uncalcined and calcined TiO2 since the intensity was low. The SEM images of LDH, TiO2, and uncalcined and calcined TiO2/ LDH nanocomposites are presented in Fig. 3. This figure shows that the LDH has a hexagonal platelet shape, with lateral dimensions of around 100 nm (Fig. 3a). Previous research reported that the morphology of LDH demonstrated a hexagonal structure due to the formation of brucite-like sheets and layer stacking (Goh et al., 2008). The morphological characteristics of TiO2 include agglomerated globular particles having an average size of approximately 30 nme50 nm (Fig. 3b). The uncalcined (Fig. 3c) and calcined TiO2/ LDH nanocomposites (Fig. 3d) revealed that TiO2 nanoparticles are attached on the lateral surface of LDH; both uncalcined and calcined TiO2/LDH nanocomposites complexion also demonstrated no morphological differences. The TiO2 nanoparticles were regularly dispersed onto the surface of LDH and successively combined with TiO2/LDH nanocomposites. DLS data revealed that the hydrodynamic size of the well-dispersed LDH, TiO2, and TiO2/LDH nanocomposites were 134.4 nm, 213.8 nm, and 322.8 nm, respectively. To separate the TiO2 nanoparticles from the calcined TiO2/ LDH nanocomposites, the sonication method used by the National Institute of Standards and Technology (NIST) was applied (Taurozzi et al., 2012); however, no individual smaller particles were observed from the well-synthesized calcined TiO2/LDH nanocomposites. Subsequently, in terms of physical interaction, it was deemed that the TiO2 nanoparticles stably combined with the LDH. To characterize the chemical interactions between TiO2 and LDH in TiO2/LDH nanocomposites, an XPS analysis was conducted. Fig. 4 shows the XPS spectra of LDH, TiO2, and uncalcined and calcined TiO2/LDH nanocomposites, with chemical states of Ti and O also presented. The spectra of the Ti 2p1/2 peak shifted from 458.8 eV in anatase TiO2 and uncalcined TiO2/LDH nanocomposites to 457.1 eV in calcined TiO2/LDH nanocomposites. The shifts of Ti peaks in calcined TiO2/LDH nanocomposites are comparable to TiO2, though there are obvious chemical differences between TiO2 nanoparticles and TiO2/ LDH nanocomposites. The TiO2 nanoparticles may represent not only the physical combination but also the chemical interaction on the calcined TiO2/LDH nanocomposites incurred by thermal treatment. Although the Ti spectra of uncalcined TiO2/LDH nanocomposites did not show a significant shift, the individual Ti spectra of uncalcined TiO2/LDH nanocomposites revealed the chemical interaction between TiO2 and LDH. Previous research also demonstrated that the individual spectra correspond to H-bonds and TieOeMe bonds (Huang et al., 2013), and that these can influence the chemical bonds between TiO2 nanoparticles and LDH nanocomposites. In addition, the differences between Ti2p1/2 and Ti2p3/2 in anatase TiO2, and uncalcined and calcined TiO2/LDH nanocomposites are the equivalent at 5.8 eV, indicating the TiO2 nanoparticles still have an octahedrally synchronized anatase structure. Fig. 4b shows the O1s peak in TiO2, and uncalcined and calcined TiO2/LDH nanocomposites, with peaks corresponding to the binding energy at 530.1 eV, 532.1 eV, and 532.1 eV, respectively. The O 1s spectra of uncalcined TiO2/LDH nanocomposites shifted 2 eV from that of anatase TiO2. Huang et al. (Huang et al., 2013) reported a similar chemical interaction between TiO2 and LDH in nano-TiO2SDS-LDHs nanocomposites, in which they suggested that the

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Fig. 3. Scanning electron microscope images of (a) LDH; (b) TiO2; (c) uncalcined TiO2/LDH (20 wt.% TiO2); (d) calcined TiO2/LDH (20 wt.% TiO2).

electron transfer may involve the reaction between TiO2 and LDH due to the positively charged LDH sheets. The two individual peaks in O1s also explain the changes in electrons at the surface of the TiO2/LDH nanocomposites due to the shifted position. The O1s spectra of the calcined TiO2/LDH nanocomposites did not significantly differ from the uncalcined TiO2/LDH nanocomposites; therefore, the chemical interaction was completed without requiring a thermal treatment. In summary, the TiO2 and LDH include not only a physical complexation, but the chemical interaction of Ti and metal complexes which involves the formation of TiO2/LDH nanocomposites. 3.2. Simultaneous photooxidation and removal process of arsenic by calcined TiO2/LDH nanocomposites Simultaneous As(III) oxidation and removal experiments were conducted using LDH, and uncalcined and calcined TiO2/LDH nanocomposites. Fig. 5a presents the batch experiment results of arsenic removal, in which LDH displays a removal capacity for As(III), As(V) and As(III) under UV irradiation of 29.80, 101.32, and 27.29 mg/g, respectively. The removal capacity of As(V) and As(III) by LDH is about 100 and 30 mg/g, respectively; this capacity was similar level to previous research (Violante et al., 2009). The data revealed that the removal capacity of As(V) is much higher than that of As(III), and that As(III) was not oxidized under UV irradiation, i.e., that the LDH does not solely induce the photooxidation of As(III). To compare the arsenic removal capacity of uncalcined and calcined TiO2/LDH, batch removal experiments was conducted using different ratios (10e30 wt.%) of TiO2 in TiO2/LDH nanocomposites. In the uncalcined TiO2/LDH nanocomposites, all samples demonstrated a lower removal capacity (<30 mg/g) than calcined TiO2/LDH. Fig 5b presents the ratios of removed and oxidized As(III) concentrations in the uncalcined and calcined TiO2/ LDH nanocomposites. The data shows that the As(III) in both

residual solutions is fully oxidized by UV irradiation. As such, the difference of removal capacity between uncalcined and calcined TiO2/LDH nanocomposites can be explained, as in Fig. 1 it was shown that the calcination effect of LDH enhances the arsenic removal capacity by increasing the potential space in the interlayer to adsorb arsenic ions (Goh et al., 2009). Fig. 6 shows the kinetic removal of As(III) and As(V) using calcined TiO2/LDH nanocomposites for different pH conditions (3, 7, and 10) under continuous UV irradiation. The optimum ratio of TiO2 in the TiO2/LDH nanocomposites was selected to be 20%, as both the removal capacity and removal rate is faster than in the 10% and 30% TiO2/LDH nanocomposites. The data showed that the removal rate of As(V) is relatively faster than that of As(III) in the TiO2/LDH nanocomposites, as the negative As(V) species is generally more favorable to the positive charged calcined LDH and due to the fact that the smaller ionic radii of As(V) (0.47 Å) is more favorable than that of As(III) (0.58 Å) to the interlayer space (7.65 Å in Fig. 1) of LDH (Kumar et al., 2014). Therefore, the As(III) removal rate under UV irradiation is slightly faster than that of As(III) in dark conditions because As(III) is continuously oxidized to As(V) which more favorably combines with LDH. In As(III) removal under dark conditions, the removal rate is not significantly affected by pH in the reaction between As(III) with TiO2/ LDH nanocomposites. However, the removal rate of As(III) in alkaline conditions (pH 10) is slightly faster than the acidic and neutral pH conditions (pH 3, 7), which may be caused by the pKa value of As(III) (H3AsO3 and H2 AsO3  ; pKa: 9.2) because alkaline pH conditions might be more favorable for the removal of As(III) due to the surface charge of As(III) species. Therefore, the As(V) removal rate decreased under alkaline conditions due to the decrease of surface charge on TiO2/LDH nanocomposites (point of zero charge: 9.8). In terms of As(III) removal under UV irradiation, the data revealed that the As(III) removal rate under alkaline conditions is faster than for acidic and neutral conditions, since

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Fig. 4. XPS spectra of anatase TiO2, uncalcined and calcined TiO2/LDH nanocomposites (a: Ti 2p, b: O 1s).

the entire reaction depends on the oxidation process; i.e., the photooxidation process is faster under alkaline conditions due to the abundance of OH, which can be converted to OH radicals. Therefore, although the overall reaction is faster than that of As(III), there is different removal pattern under various pH conditions compared with the results of As(V).

3.3. Photooxidation mechanism of As(III) by calcined TiO2/LDH nanocomposites The TiO2-catalyzed photooxidation of arsenite can be induced by the generation of positive holes ðhvb þ Þ, electrons ðe cb Þ, hydroxyl radicals (OH), superoxides (O2), and hydrogen peroxide

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Fig. 5. (a) arsenic removal by LDH and TiO2/LDH nanocomposites (As(III) only, As(V) only, As(III) with UV irradiation) (b) Oxidation ratio of As(III) in react with TiO2/LDH nanocomposites under UV irradiation (initial concentration of arsenic: 50 mg/l; dose: 0.5 g/l; reaction time: 24hr; temperature: 30  C; 180 rpm).

Fig. 6. Simultaneous As(III) photooxidation and removal by TiO2/LDH nanocomposites (20%) under UV irradiation.

(Table 1a). Indeed, the photooxidation mechanism of As(III) by TiO2 nanoparticles has posed a controversial problem. Many researches have reported that superoxides and hydroxyl radicals are not the main oxidants during As(III) oxidation, even though hydroxyl radicals have been a key factor to photodegrade organic and inorganic contaminants in drinking water (Yoon and Lee, 2005). For example, Choi et al. (Ryu and Choi, 2006) reported that superoxide-mediated photooxidation process is the main oxidant in the As(III) oxidation process, and Yoon et al. (Yoon et al., 2009) have claimed that the superoxide has little effect on the photooxidation of As(III), whereas As(III) can be initially oxidized to As(IV) by hvb þ, OHad and then finally converted to As(V) by oxygen, hvb þ and OHad. In order to further investigate the photooxidation mechanism of As(III) by TiO2/LDH nanocomposites, excessive concentrations of tert-BuOH (50 mM) and formic acid (10 mM) were added into the As(III) removal experiments of this study (Fig. 7a) because the tertBuOH is an OH radical scavenger and formic acid is a valence band hole scavenger (Xu et al., 2005; Yoon et al., 2009). Fig. 7 shows that

As(III) can be removed by the addition of TiO2/LDH nanocomposites in an ion exchange process (black circles), whereas the TiO2/LDH nanocomposites generated oxidants such as superoxides and hydroxyl radicals under UV irradiation and that the removal rate also increased with the continuous oxidation of As(III) (white circles). However, the addition of formic acid (white triangles) indicates a negative effect for the removal of As(III) with UV irradiation, as the rate was even slower than that of TiO2/LDH with no UV irradiation. This result can be explained by the fact that the inhibition of As(III) photooxidation by supported TiO2 particles can be originated from the effect of the valence band inhibition. As shown in Table 1(c), formic acid generates strong reducing radicals, and As(IV) can be rapidly reduced by the effect of reducing radicals even though As(III) is oxidized by hvb þ din fact, As(III) was the dominant species in the final solution in the formic acid addition samples (Choi et al., 2010; Yoon et al., 2009). Furthermore, the second effect is the inhibition of the ion exchange process due to the excessive formic acid as formic acid covers the surface of TiO2/LDH nanocomposites

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Table 1 Photooxidation process of As(III) by supported TiO2. Ref. a. General photocatalytic reaction of supported TiO2 Supported TiO2 þ hv/e CB þ hþ VB e CB þ hþ VB /heat OHads þ hþ VB /  OH O2 þ e CB /O2   O2   þHþ /HO2  O2   þe CB þ Hþ /H2 O2 O2   þO2   þ2Hþ /H2 O2 þ O2 H2 O2 þ e CB /  OH þ OH b. Photooxidation process of As(III) As(III) þ OH / As(IV) þ OHAsðIIIÞ þ hþ VB ðor surface  bound OHÞ/AsðIVÞ AsðIIIÞ þ O2 þ Hþ /AsðIVÞ þ HO2  As(IV) þ O2 / As(V) þ O2/HO2 c. Inhibition of As(III) photooxidation t-BuOH (OH radical scavenger) OH þ t-BuOH / H2O þ CH2C(CH3)2OH formic acid (hþ vb scavenger) HCO2 H þ hþ vb /HCO2  þHþ 2HCO2 þ As(V) / CO2 þ As(III) þ 2Hþ HCO2 þ As(IV) / CO2 þ As(III) þ Hþ

k ¼ 2.0  1010 M1s1 pKa ¼ 4.8

Dutta et al., 2005 Rush and Bielski, 1985

k ¼ 8.3  105 M1s1

Rush and Bielski, 1985

k ¼ 9.0  109 M1s1 k ¼ 3.6  10 M s k ¼ 1.1  109 M1s1

Ryu and Choi 2006 Yoon et al., 2005 Ryu and Choi 2006 Ryu and Choi 2006

k ¼ 6.0  108 M1s1

Xu et al., 2005

6

1 1

Yoon et al., 2009 Choi et al., 2010

Fig. 7. (a) As(III) removal by TiO2/LDH nanocomposites with tert-buthanol and formic acid (initial concentration of As(III): 10 mg/l, initial pH: 3, dose: 0.5 g/L) (b) Normalized As Kedge XANES spectra recorded for As(III) and As(V) adsorbed on calcined TiO2/LDH with UV irradiation.

and this effect can then inhibit the ion exchange process of As(III) in TiO2/LDH nanocomposites. In contrast, excessive t-BuOH did not inhibit the removal process of As(III) even though free OH radicals were generated by UV irradiation. This finding can be explained by the fact that the addition of OH radical scavengers was not significant in the photooxidation process, and therefore the OH radicals were not the main oxidant of the As(III) photooxidation process. To determine the oxidized arsenic species on the surface of TiO2/ LDH nanocomposites, the XANES study was conducted. Fig. 7b presents the arsenic K-edge XANES spectra of TiO2/LDH nanocomposites for As(III), As(III) under UV irradiation for 6 h and As(V). Here, NaAs(III)O2 and NaH2As(V)O4 were used as standard samples for As(III) and As(V), respectively (Takahashi et al., 2004). The XANES spectra of TiO2/LDH shows that the UV irradiation clearly oxidizes As(III) to As(V) on the surface of TiO2/LDH nanocomposites. The standard spectrum of As(III) is located at 11,866 eV, and closely corresponds to that of As(III) in the TiO2/LDH nanocomposites sample. The calculated ratios of As(III) and As(V) in the samples were 70.7% and 29.3%, respectively; this wide

spectrum can be explained by the partial oxidation of As(III) to As(V) during sample preparation. The standard As(V) spectra is positioned at 11,868 eV, and the obtained spectrum fully matches the spectra of As(V) and As(III) with UV irradiation samples. The calculated ratios of As(III) and As(V) in these samples were almost 99%. Therefore, this results clearly confirms that As(III) was successfully oxidized by UV irradiation and that the oxidized species of arsenic combined with the TiO2/LDH nanocomposites. 4. Conclusion Simultaneous photooxidation and ion exchange process were successively implemented as bi-functional material for As(III) removal in water. The synthesized calcined TiO2/LDH nanocomposites had both the crystallographic properties of anatase TiO2 and calcined LDH. The SEM data revealed that the TiO2 nanoparticles (<50 nm) were attached on the LDH surface (around 100 nm). The XPS data further revealed that TiO2 nanoparticles were chemically combined with the LDH, and the DLS data showed the

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TiO2/LDH nanocomposites were stably present as monoparticles in the water. The calcination process of the TiO2/LDH nanocomposites enhanced the removal capacity by increasing interlayer space used to implement the ion exchange process. In these experimental conditions, 20% TiO2 loaded TiO2/LDH nanocomposites displayed the advantages for both removal capacity and rate. The simultaneous As(III) photooxidation and removal process were more favorable in alkaline conditions than acidic conditions, and the direct oxidation of hvb þ was deemed to be the main role of As(III) oxidation in the process. Subsequent XANES data confirmed that the combined arsenic species is dominantly As(V); therefore, this novel photooxidation and removal process was successively applied to the simultaneous photooxidation and removal process of As(III) in water. Acknowledgement This work was supported by the BK21 plus program through the National Research Foundation (NRF) funded by the Ministry of Education of Korea (No. 21A20130012512) References Anawar, H.M., Akai, J., Mostofa, K.M.G., Safiullah, S., Tareq, S.M., 2001. Arsenic poisoning in groundwater: health risk and geochemical sources in Bangladesh. Environ. Int. 27, 597e604. Bissen, M., Frimmel, F.H., 2003. Arsenic - a review. Part I: occurrence, toxicity, speciation, mobility. Acta Hydrochim. Hydrobiol. 31, 9e18. Chapagain, S.K., Shrestha, S., Nakamura, T., Pandey, V.P., Kazama, F., 2009. Arsenic occurrence in groundwater of Kathmandu Valley, Nepal. Desalin. Water Treat. 4, 248e254. Cheng, X., Huang, X., Wang, X., Zhao, B., Chen, A., Sun, D., 2009. Phosphate adsorption from sewage sludge filtrate using zinc-aluminum layered double hydroxides. J. Hazard. Mater. 169, 958e964. Cho, S., Kwag, J., Jeong, S., Baek, Y., Kim, S., 2013. Highly fluorescent and stable quantum dot-polymer-layered double hydroxide composites. Chem. Mater. 25, 1071e1077. Choi, W., Yeo, J., Ryu, J., Tachikawa, T., Majima, T., 2010. Photocatalytic oxidation mechanism of As(III) on TiO2: unique role of As(III) as a charge recombinant species. Environ. Sci. Technol. 44, 9099e9104. Das, N.N., Konar, J., Mohanta, M.K., Srivastava, S.C., 2004. Adsorption of Cr(VI) and Se(IV) from their aqueous solutions onto Zr4þ-substituted ZnAl/MgAl-layered double hydroxides: effect of Zr4þ substitution in the layer. J. Colloid Interface Sci. 270, 1e8. lica, R.S., 2007. Arsenic occurrence in Brazil and De Figueiredo, B.R., Borba, R.P., Ange human exposure. Environ. Geochem. Health 29, 109e118. Dutta, P.K., Phekonen, S.O., Sharma, V.K., Ray, A.K., 2005. Photocatalytic oxidation of arsenic(III): evidence of hydroxyl radicals. Environ. Sci. Technol. 39, 1827e1834. Environmental Protection Agency, 2001. National Primary Drinking Water Regulation: arsenic and clarifications to compliance and new source contaminants monitoring: final rule. Fed. Regist. 66, 6976e7066. Goh, K.-H., Lim, T.-T., Dong, Z., 2009. Enhanced arsenic removal by hydrothermally treated nanocrystalline Mg/Al layered double hydroxide with nitrate intercalation. Environ. Sci. Technol. 43, 2537e2543. Goh, K.-H., Lim, T.-T., Dong, Z., 2008. Application of layered double hydroxides for removal of oxyanions: a review. Water Res. 42, 1343e1368. Huang, Z., Wu, P., Lu, Y., Wang, X., Zhu, N., Dang, Z., 2013. Enhancement of photocatalytic degradation of dimethyl phthalate with nano-TiO2 immobilized onto hydrophobic layered double hydroxides: a mechanism study. J. Hazard. Mater. 246e247, 70e78. Jiao, C., Chen, X., Zhang, J., 2009. Synergistic effects of Fe2O3 with layered double hydroxides in EVA/LDH composites. J. Fire Sci. 27, 465e479. Kameda, T., Nakamura, M., Yoshioka, T., 2012. Removal of antimonate ions from an aqueous solution by anion exchange with magnesium-aluminum layered double hydroxide and the formation of a brandholzite-like structure. J. Environ. Sci. Health. A. Tox. Hazard. Subst. Environ. Eng. 47, 1146e1151. €, M., 2014. Interaction Kumar, E., Bhatnagar, A., Hogland, W., Marques, M., Sillanp€ aa of anionic pollutants with Al-based adsorbents in aqueous media - a review. Chem. Eng. J. 241, 443e456. Kurttio, P., Pukkala, E., Kahelin, H., Auvinen, A., Pekkanen, J., 1999. Arsenic concentrations in well water and risk of bladder and kidney cancer in Finland. Environ. Health Perspect. 107, 705e710. Li, H., Zhu, G., Liu, Z.H., Yang, Z., Wang, Z., 2010. Fabrication of a hybrid graphene/ layered double hydroxide material. Carbon N. Y. 48, 4391e4396. Li, Y., Cai, X., Guo, J., Na, P., 2014. UV-induced photoactive adsorption mechanism of arsenite by anatase TiO2 with high surface hydroxyl group density. Colloids Surf. A 462, 202e210.

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