Efficient Photocatalytic Decomposition of Perfluorooctanoic Acid by

Apr 10, 2012 - ABSTRACT: Perfluorooctanoic acid (C7F15COOH, PFOA) has increasingly attracted worldwide concerns due to its global occurrence and ...
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Efficient Photocatalytic Decomposition of Perfluorooctanoic Acid by Indium Oxide and Its Mechanism Xiaoyun Li,† Pengyi Zhang,*,† Ling Jin,† Tian Shao,† Zhenmin Li,† and Junjie Cao† †

State Key Joint Laboratory of Environment Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing 100084, China S Supporting Information *

ABSTRACT: Perfluorooctanoic acid (C7F15COOH, PFOA) has increasingly attracted worldwide concerns due to its global occurrence and resistance to most conventional treatment processes. Though TiO2-based photocatalysis is strong enough to decompose most organics, it is not effective for PFOA decomposition. We first find that indium oxide (In2O3) possesses significant activity for PFOA decomposition under UV irradiation, with the rate constant about 8.4 times higher than that by TiO2. The major intermediates of PFOA were C2−C7 shorter-chain perfluorocarboxylic acids, implying that the reaction proceeded in a stepwise manner. By using diffuse reflectance infrared Fourier transform spectroscopy, 19F magic angle spinning nuclear magnetic resonance, and electron spin resonance, we demonstrate that the terminal carboxylate group of PFOA molecule tightly coordinates to the In2O3 surface in a bidentate or bridging configuration, which is beneficial for PFOA to be directly decomposed by photogenerated holes of In2O3 under UV irradiation, while PFOA coordinates to TiO2 in a monodentate mode, and photogenerated holes of TiO2 preferentially transform to hydroxyl radicals, which are inert to react with PFOA. PFOA decomposition in wastewater was inhibited by bicarbonate and other organic matters; however, their adverse impacts can be mostly avoided via pH adjustment and ozone addition.

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INTRODUCTION Perfluorooctanoic acid (PFOA), as a representative of perfluorinated chemicals (PFCs), has been detected globally in water, air, wildlife, and humans. PFOA has some particular properties such as extraordinary thermal, chemical stability, and high surface-active effect owing to the C−F bond structure, which makes it widely used as surfactants and fire retardants in the past decades.1,2 In recent years, many studies indicate that PFOA is environmentally persistent and bioaccumulative.3,4 Furthermore, its toxicity and likely carcinogenic effect on human and wildlife have been proved.5,6 In view of these facts, PFOA has been restricted to use in some countries and regions. In addition, researchers are making efforts to develop costeffective treatment methods to eliminate PFOA from environment. It has been reported that photochemical and sonochemical methods are two types of treatment technologies to degrade PFOA efficiently.7−17 Hori et al. reported that sulfate radical anions (SO4·‑) produced by S2O82‑ photolysis can decompose PFOA and other shorter-chain perfluorocarboxylic acids (PFCAs) via an electron transfer from PFOA to SO4·‑ radical.7,8 In other oxidative processes, a similar decomposition mechanism was proposed, i.e. PFOA first loses an electron and then it is decarboxylated to form a perfluoroheptyl radical.9−11 On the other hand, PFOA can be reductively decomposed by aqueous electrons (eaq−) via a defluorination step.12 As a typical semiconductor photocatalyst, TiO2 has been extensively studied to eliminate various organic pollutants. © 2012 American Chemical Society

However, it has been reported that TiO2 has low activity to decompose PFOA because PFOA is almost inert to hydroxyl radicals (HO·) generated in the TiO2 photocatalysis process (kOH·+PFOA ≤ 105 M−1 s−1).18,19 PFOA contains no hydrogen atoms available for abstraction by HO·. Perfluorination reduces electron density of −COO− group in PFOA, thus the direct electron transfer between HO· and −COO− group in PFOA is not favorable.19 Similarly, TiO2 shows low activity for trichloroacetate, which does not have C−H bond either.20−22 Only under extreme conditions, for example, in a highly acidic aqueous solution (HClO4, pH < 2), the activity of TiO2 for PFOA and other shorter-chain PFCAs is enhanced.23,24 On the other hand, some other semiconductor materials show better activity and stability than TiO2 for photocatalytic degradation of some persistent pollutants.25,26 After screening several semiconductor materials such as ZnS, NiO, and Ga2O3, we found that In2O3 possessed higher activity than that of TiO2 for PFOA decomposition. In the present study, we report the heterogeneous photocatalytic decomposition of PFOA by In2O3 and elucidate why In2O3 shows better activity for PFOA decomposition than TiO2. Received: Revised: Accepted: Published: 5528

November 29, 2011 April 5, 2012 April 10, 2012 April 10, 2012 dx.doi.org/10.1021/es204279u | Environ. Sci. Technol. 2012, 46, 5528−5534

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EXPERIMENTAL SECTION Materials. Indium oxide (In2O3, 99.99%; BET surface area, 12.6 m2/g) was obtained from Sinopharm Chemical Reagent Co. Ltd. (China). Its crystal structure was pure cubic phase (JCPDS No. 71-2194) as characterized by XRD (Figure S1 in the Supporting Information). TiO2 (Degussa P25; BET surface area, 50 m2/g; antase/rutile ratio, 80:20) was used as the reference photocatalyst. PFOA (96%, Acros), perfluoroheptanoic acid (PFHpA, 96%, Acros), perfluorohexanoic acid (PFHxA, 97%, Fluka), perfluoropentanoic acid (PFPeA, 97%, Acros), perfluorobutanoic acid (PFBA, 99%, Acros), perfluoropropanoic acid (PFPrA, 97%, Acros), and trifluoroacetic acid (TFA, 99%, Acros) were used with received. A spin-trap reagent, i.e. 5,5-dimethyl-1-pyrroline-N-oxide (DMPO), was acquired from Tokyo Chemical Industry Co., Ltd. (Japan). All aqueous solutions were prepared with high purity water (18.2 MΩ) prepared by using the Thermo Barnstead Nanopure Diamond water purification system. Photocatalytic Procedures. The photocatalytic experiments were conducted in a tubular glass reactor with the inner diameter of 45 mm. A low-pressure mercury lamp (23 w, Cnlight Co. Ltd., Guangdong, China) emitting 254 nm UV light was placed in the center of the reactor with a quartz envelope. Concentrations of PFCs in the natural waters are generally in the range of pg/L to ng/L, but high concentrations on the order of mg/L have been measured in the receiving waters near specific point sources and in some industrial wastewaters.27,28 Considering the relative high adsorbability of In2O3 and water solubility of PFOA (3.4 g/L),29 initial concentration of PFOA as high as 100 μmol/L was used in the experiments to reduce the impact of In2O3 adsorption on the PFOA measurement. In a typical experiment, an aqueous PFOA solution (400 mL, 100 μmol/L) was mixed with 0.2 g of photocatalyst (In2O3 or TiO2) in a beaker followed by stirring to obtain a uniform suspension. Samples before and after mixing were taken to learn the amount of PFOA adsorbed by photocatalysts. Then the suspension was poured into the reactor and purged with oxygen gas for 30 min prior to UV irradiation. The reaction temperature was kept at 25 °C with a cooling water jacket around the reactor, and oxygen gas was continuously bubbled into the reactor all the time. Aliquots of sample were taken at a regular time interval and filtrated with 0.22-μm filters to remove photocatalyst for subsequent analysis. To avoid possible impact of filter adsorption on PFOA concentration, the first 5 mL filtrate was discarded. The gasphase intermediates were collected for 20 min at regular time intervals with a thermal desorption tube (Carbosieve SIII, Supelco). Analysis. Concentrations of PFOA and its decomposition products were measured on a Waters Acquity UPLC system, coupled with a Micromass Quattro Premier tandem quadrupole mass spectrometric system (Waters, Milford, USA). Prior to analysis, each sample was diluted 50 times to ensure the PFOA concentration below 1 mg/L, and 9 aqueous standard solutions containing PFOA and C2−C7 shorter-chain PFCAs were prepared to make external calibration curves in the range of 2−900 μg/L. The multiple reaction monitor mode (MRM) was used for quantitative analysis of PFOA and other shorter-chain PFCAs. The separation column was a Waters Acquity UPLC BEH C18 column (2.1 mm i.d. × 50 mm, 1.7 μm particles), and the column temperature was set at 50 °C. The flow rate was maintained at 0.4 mL/min with a mobile phase of eluent A (2

mmol/L ammonium acetate/methanol) and B (2 mmol/L ammonium acetate/water). The eluent gradient started with 30% A for 0.5 min, then was linearly increased to 90% A in 4.5 min, and further increased to 100% A in 1 min and finally back to 30% A in 3 min. MS detection was operated in negative mode by using an electrospray source. The optimum mass parameters obtained were as follows: capillary voltage, 2.1 kV; source temperature, 120 °C; desolvation temperature, 280 °C; and desolvation gas flow rate, 650 L/h. The parent ion, daughter ion, cone voltage (V), and collision energy (eV) for PFCAs detection were PFOA (413 > 369, 17, 11), PFHpA (363 > 319, 16, 10), PFHxA (313 > 269, 15, 10), PFPeA (263 > 219, 15, 10), PFBA (213 > 169, 12, 10), PFPrA (163 > 119, 12, 10), and TFA (113 > 69, 12, 10), respectively. In addition, the full scan mode was used to detect parent ions for acquiring more information about intermediates. Samples were directly infused into the mass spectrometer ion source by pure methanol with the cone voltage of 10 V. The concentration of F− was measured with an ion chromatography system (Dionex ICS-2000, USA) consisting of a degasser, an autosampler (250 μL injection volume), a guard column (IonPac AG11-HC, 4 × 50 mm), a separation column (IonPac AS11-HC, 4 × 250 mm), a column heater (30 °C), and a conductivity detector with a suppressor. The mobile phase was an aqueous solution of KOH (30 mmol/L), and the flow rate was 1.0 mL/min. The suppressor current was set at 124 mA. The qualitative analysis of gas-phase intermediate was conducted on a GC/MS (GCMS-QP2010 Plus, SHIMADZU, Japan). Prior to the GC/MS analysis, the adsorbed gases were desorbed from the adsorption tube by an automatic thermal desorber (Turbo Matrix ATD 650, PerkinElmer). Characterization of Photocatalysts. The fresh and PFOA-adsorbed photocatalysts (In2O3 and TiO2) were characterized by BET surface area, X-ray diffraction (XRD), X-ray photoelectron spectroscopy (XPS), diffuse reflectance infrared Fourier transform spectroscopy (DRIFTS), 19F magic angle spinning nuclear magnetic resonance (19F MAS NMR), and electron spin resonance (ESR). BET-surface area of the commercial In2O3 powder was measured by nitrogen adsorption at −196 °C, using a Quantachrome autosorb-1 apparatus. XRD measurements was carried out on a Rigaku D/Max 2500 diffractometer with a Cu kα monochromatized radiation source, operated at 40 kV and 200 mA. The photocatalyst powder was filtered and dried at room temperature for 24 h before XPS measurement. XPS analysis was conducted on a PHI 5300 ESCA instrument using Al Ka as an exciting X-ray source. The pass energy of the analyzer was set at 55 eV. The spectra were calibrated with respect to the C1s line of adventitious carbon at 284.8 eV. DRIFTS measurements were carried out on a Nicolet NEXUS 670-FTIR spectrometer equipped with a smart collector and an MCT detector cooled by liquid N2. The IR spectra were recorded by accumulating 100 scans at a resolution of 4 cm−1. The 19F MAS NMR experiments were performed at room temperature on a Bruker Avance III 400 MHz spectrometer equipped with a 2.5 mm Bruker 1H/19F/X probe, operating at 376.20 MHz for 19F. The powder samples were loaded into a zirconia rotor and spun at an MAS rate (vr) of 25 kHz. The 19F π/2 pulse width was 3.0 μs, the relaxation delay was 3 s, and the acquisition time was 18 ms. Numbers of scans acquired for PFOA with KBr, PFOA adsorbed on In2O3 or TiO2 were 256, 4000, and 8000, respectively. The peak positions were 5529

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ability than TiO2 to adsorb PFOA. Besides the decrease of PFOA concentration, the pH value increased from 3.8 to 4.2 after addition of In2O3, which implied that the surface hydroxyl group of In2O3 was replaced by PFOA and released into the aqueous phase. After irradiated with 254 nm UV light, a slight direct photolytic decomposition of PFOA (9.8% after 4 h) was observed, which is in agreement with reports in the literature.9,31 In the presence of TiO2, 15.9% PFOA was decomposed within 4 h, which is about two times higher than that by the direct photolysis. While in the presence of In2O3 material, the decomposition ratio of PFOA dramatically increased to over 80% after 4 h irradiation. The pH value also decreased to 3.1 due to formation of acidic products, such as HF and organic intermediates. As shown in Figure S2 in the Supporting Information, the PFOA decomposition under different conditions followed pseudofirst-order kinetics with the rate constants of 0.378 h−1 for In2O3, 0.045 h−1 for TiO2, and 0.026 h−1 for the direct photolysis, respectively. Thus, the decomposition rate constant of PFOA by In2O3 is 8.4 times higher than that by TiO2. The main decomposition products of PFOA are C2−C7 shorter-chain PFCAs and fluoride ion, of which C2−C7 shorterchain PFCAs were quantified by UPLC/MS/MS and fluoride ion was measured by ion chromatography. The chromatogram spectra and the calibration curves of PFOA and C2−C7 shorterchain PFCAs products for UPLC/MS/MS are shown in Figure S3 and S4 in the Supporting Information. The time dependence of the formation of C2−C7 shorter-chain PFCAs and fluoride ion in the presence of In2O3 are shown in Figure 1(b). The amount of PFHpA increased rapidly at the beginning, reached a maximum concentration in 3 h, and then gradually decreased, while other PFCAs bearing shorter perfluoroalkyl groups increased continuously during the 4 h reaction period. It indicates that PFOA is decomposed in a stepwise manner to remove the CF2 unit one by one. The amount of F− increased with increase of irradiation time. The mass balance of fluorine element (F) in the reaction system was calculated in detail, and the results are summarized in Table 1. Total F content in aqueous solution consists of four parts, i.e. remaining PFOA, shorter-chain PFCAs, F− and PFCAs adsorbed on catalyst surface, was kept at about 88%. The F content on the catalyst surface was estimated in terms of the adsorption amount and the XPS quantitative result (Table S1 in the Supporting Information). The other 12% F content may be transformed to the gas-phase products. As analyzed by ATD/GC-MS, a small amount of HCOOH and C4−C6 alkane such as CnF2n+2 and CnHF2n+1 were detected in the gas phase. No other intermediates are detected in the aqueous solution after 4 h reaction besides C2−C7 shorter-chain PFCAs by fullscan mode of UPLC/MS/MS analysis. Coordination of PFOA to In2O3 and TiO2. Adsorption of reactants to the photocatalyst surface is a critical step in the photocatalytic process.32 As reported by Moser et al.,33 surface coordination of TiO2 by benzene derivatives enhanced the interfacial electron transfer rate, and bidentate derivatives showed much higher enhancement than monodentate benzoate. The DRIFTS is a highly sensitive method to characterize structural changes of adsorbed species, and the spectra of PFOA adsorbed on In2O3, TiO2, and mixed with KBr powder are shown in Figure 2. An absorbance at 1769 cm−1 in IR spectrum of mixture of PFOA and KBr is typical of CO vibration for carboxylic acid. The strong bands in the range of

referenced to the amorphous domains in poly(vinylidene fluoride) at δF = −88.7 ppm.30 ESR signals of radicals trapped by DMPO were recorded on a JEOL JES FA-200 (Japan) spectrometer at ambient temperature under photoirradiation of a mercury lamp (ESUSH500, 500W) with a filter (UV-D35, 76% transmission for 270−440 nm, 24% transmission for >680 nm). Typical spectrometer parameters were as follows: center field 3378 G, sweep width 100 G, microwave frequency 9.440 GHz, modulation frequency 100 kHz, power 1 mW. The magnetic field was calibrated with a Mn2+ marker. In order to quantitatively compare the concentration of hydroxyl radical under different conditions, the amount of solution sucked in the quartz capillary tube and the photoirradiation time were the same for all ESR measurements. Other conditions were as follows: the pH value of pure aqueous solution, 3.8 (adjusted with H3PO4); PFOA concentration, 100 μmol/L (pH≈3.8); the photocatalyst dosage, 0.5 g/L; and DMPO concentration, 0.04 mol/L.

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RESULTS AND DISCUSSION Photocatalytic Decomposition of PFOA by In2O3 and TiO2. Figure 1(a) shows the adsorption and photocatalytic decomposition of PFOA by In2O3 and TiO2. After In2O3 and TiO2 powders were added to the PFOA aqueous solution, PFOA concentration decreased 15.0 and 7.1 μmol/L, respectively. This indicates that In2O3 has much stronger

Figure 1. (a) Time course of PFOA decomposition under UV irradiation in the presence of In2O3, TiO2, or without a photocatalyst and (b) time course of shorter-chain intermediates in the presence of In2O3. Each error bar represents one standard deviation from the mean of at least three experiments. 5530

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Table 1. F and C Content in the Reactant and Products after PFOA Decomposition in the Presence of In2O3a F content in PFOA and products (%) reaction time (h)

remaining PFOA

shorter chains

F−

cat. surface

2

44.9

26.9

15.3

ndc

4

16.9

36.4

33.7

1.4

d

gas-phase productsb C4: C4F10, C4F9H; C5: C5F12, C5F11H; C6: C6F14, C6F13H; HCOOH, CH3COOH

total F content in liquid phase (%)

total carbon content in liquid phase (%)

87.1

72.7

88.4

54.8

a

The reaction conditions were the same as in Figure 1. bGas-phase products were collected with a thermal desorption tube and qualitatively analyzed by GC-MS. cnd = not detected. dThe calculated method was described in Table S1.

greater than that of K+ salt (203 cm−1), which indicates that PFOA coordinates to TiO2 in a unidentate mode. While the Δν value for PFOA/In2O3 (198 cm−1) is somewhat smaller than ionic Δν, considering the stronger adsorption of PFOA on In2O3 (In2O3, 30 μmol/g; TiO2, 14.2 μmol/g), it can be concluded that PFOA is bound to In2O3 in a bidentate or bridging configuration. The 19F MAS NMR spectra of PFOA and samples adsorbed on In2O3 or TiO2 are shown in Figure 3. For the sample of

Figure 2. DRIFT spectra of PFOA mixed with KBr and adsorptionequilibrium on In2O3 and TiO2 photocatalysts at room temperature. The samples of PFOA/photocatalyst were filtered and dried at room temperature for 24 h before measurement.

Figure 3. 19F MAS NMR spectra of PFOA mixed with KBr and adsorption-equilibrium on In2O3 and TiO2 photocatalysts at room temperature. The samples of PFOA/photocatalyst were filtered and dried at room temperature for 24 h before measurement.

1300−1100 cm−1 are assigned to C−F stretching. The C−OH vibration is overlaid with an absorbance peak at 1210 cm−1.34 The broad bands at 1640−1600 cm−1 are assigned to adsorbed water molecules. As for potassium perfluorooctanoate (PFOK), the ν(CO) vibration peak is replaced with the asymmetric (νas(COO−)) and symmetric stretches (νs(COO−)) of carboxylate, appearing at 1660 cm−1 and 1457 cm−1, respectively. Upon adsorption on In2O3, the ν(CO) stretch of PFOA vanishes, and two new peaks appear at 1641 and 1443 cm−1, which are assigned to the asymmetric and symmetric stretching modes of the −COO group.34−37 Similarly, upon adsorption on TiO2, the ν(CO) vibration peak disappears, and νas(COO−) and νs(COO−) appear at 1686 cm−1 and 1408 cm −1 , respectively. These results indicate that PFOA coordinates with both In2O3 and TiO2 via its carboxylate group. According to the stretching frequencies of the carboxylate, Deacon et al. concluded an empirical relationship between the frequency difference,38 Δν = νas(COO−) − νs(COO−), and the types of bonding of carboxylate to cations. The Δν value which is substantially greater than the ionic Δν indicates a monodentate coordination, while the Δν value which is significantly less than the ionic Δν indicates a bidentate or bridging coordination. When the Δν value is near ionic Δν, chelating and/or bridging carboxylates cannot be excluded. The Δν value observed for PFOA/TiO2 (278 cm−1) is substantially

PFOA/KBr, peaks at −80.1, −116.9, and −124.6 ppm are assigned to terminal CF3, the CF2 group next to the carboxylic group, and the CF2 group adjacent to terminal CF3, respectively. Peaks in the range from −119.6 to −120.9 ppm are assigned to other CF2 groups. The peak assignments are consistent with the result of Buchanan et al.,39,40 who used 19 F−19F correlation spectrometry technique to confirm 19F chemical shifts of PFOA in solution. After PFOA adsorbed, changes of the chemical shifts and line widths were observed. The terminal CF3 group showed a significant shift to high field about 1.5 and 1.8 ppm for In2O3 and TiO2 respectively, and the CF2 group adjacent to CF3 group also shifted upfield. The shift of CF3 and adjacent CF2 group can be attributed to these groups located at the air/monolayer interface as Pawsey et al.41 explained for PFCAs adsorbed on ZrO2. In addition, the line width of CF2 group next to the carboxylic group broadened, which is attributed to the deprotonation of carboxylic acid group and its coordination to the photocatalyst. Notably, there is a great difference in 19F line width of other inner CF2 groups between In2O3 and TiO2. For the PFOA/TiO2, the peaks heavily overlapped and lumped into a big peak, while those for PFOA/In2O3 were less changed and remained a shape similar 5531

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to those of the bulk acid. These observations indicate that the inner CF2 groups of PFOA may interact with TiO2 surface and hardly interact with In2O3. Considering the results from DRIFTS and 19F MAS NMR, schematic configurations of PFOA adsorbed on In2O3 and TiO2 are shown in Figure 4. As for PFOA/In2O3, PFOA closely

Figure 5. DMPO spin-trapping ESR spectra under UV irradiation for 4 min at room temperature of water and PFOA solution in the presence of In2O3 or TiO2.

generated holes could directly react with PFOA and thus fewer holes were transformed into OH radicals. According to the ESR results, possible mechanisms of PFOA decomposition on In2O3 and TiO2 are summarized in Figure 6. Compared with TiO2 Figure 4. Schematic diagram of PFOA configurations adsorbed on In2O3 and TiO2.

coordinates to In2O3 in a bidentate or bridging mode, resulting in a vertical and ordered configuration of PFOA chain on In2O3 surface. While in the case of PFOA/TiO2, PFOA binds on TiO2 surface in a monodentate mode with its carboxylate group, resulting in a tilted configuration of PFOA on TiO2 surface. Consequently, the inner CF2 group of PFOA may interact with TiO2 surface OH group via hydrogen bonds. The formation of hydrogen bonds at the organic−inorganic interface have been widely reported and observed by solid-state NMR.42 Mechanism of PFOA Decomposition by In2O3. It is well-known that photocatalytic oxidation can proceed via direct hole oxidation or via indirect ·OH radicals, and the tight adsorption of the electron donor is an essential requisite for direct hole oxidation.32 As illustrated above, PFOA closely coordinates to In2O3 in a bidentate or bridging configuration, and it is reasonable that PFOA decomposition may be induced by direct hole oxidation. We measured ·OH radicals formed in the presence of different photocatalysts by ESR using DMPO as the spin-trap reagent. As shown in Figure 5, the characteristic four peaks of DMPO-·OH with intensity 1:2:2:1 appear in all spectra. The intensities of ·OH generated in TiO2 suspension were much greater than those in In2O3 suspension, which indicates that the photogenerated holes in the valence band of TiO2 are largely transformed into ·OH radicals, while those of In2O3 react slowly with surface-bound water or hydroxyl group (−OH). After addition of PFOA, the signals of ·OH peaks increased greatly in TiO2 system, which is consistent with previous reports that the F− and CF3COO− adsorbed on TiO2 could enhance the production of free ·OH radicals owing to their strong electron-withdrawing ability to reduce the recombination of electrons and holes.43−45 Because photogenerated holes of TiO2 are quickly and mostly transformed into ·OH radicals after addition of PFOA, and ·OH radicals are not effective to degrade PFOA,18,46 it is understandable that TiO2 shows a low activity for PFOA decomposition. However, intensities of ·OH radicals in In2O3 system were slightly lowered after addition of PFOA, which implied that photo-

Figure 6. Possible mechanisms of the photocatalytic decompostion of PFOA by In2O3 and TiO2.

material, In2O3 has a higher adsorption capacity and tightly coordinates with PFOA in a bidentate or bridging configuration, which is beneficial for PFOA to be decomposed via direct hole oxidation. Accordingly, In2O3 possesses much higher activity than TiO2 to decompose PFOA under UV irradiation. Potential Applications. In real wastewater, PFOA generally coexists with other chemical compounds, such as organic pollutants, natural organic matter, and bicarbonate, which may reduce the PFOA decomposition efficiency.14,17 To validate the feasibility of In2O3 photocatalysis to decompose PFOA in a real wastewater, we investigated the decomposition of PFOA added in a secondary effluent taken from a municipal wastewater plant in Beijing, China. As shown in Table S2 in the Supporting Information, the secondary effluent contained high concentration bicarbonate (4.76 mmol/L), and its TOC and pH were 18.9 mg/L and 7.8, respectively. As shown in Figure 7, the decomposition of PFOA in the original secondary effluent was almost inhibited. Moreover, the adsorption of PFOA on In2O3 became insignificant, which can be attributed to the competitive adsorption of bicarbonate anion on In2O3 carrying positive charges at pH 7.8 (the pHzpc of In2O3 is 8.747). Bicarbonate (HOCOO−) has the same carboxyl group as PFOA, and its 5532

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AUTHOR INFORMATION

Corresponding Author

*Phone: +86-10-62796840 ext. 601. Fax: +86-10-62797760. Email: [email protected]. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS This work was financially supported by National Natural Science Foundation of China (No. 21177071), Tsinghua University Initiative Scientific Research Program and special fund of State Key Joint Laboratory of Environment Simulation and Pollution Control.

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Figure 7. Photocatalytic decompostion of PFOA in pure water (PW) or in the secondary effluent (SE) by In2O3. PFOA concentration added in SE was the same as that in PW. The original pH of SE was 7.8 and was adjusted with HCl; ozone was provided with two ozonegenerating UV lamps.

(1) Key, B. D.; Howell, R. D.; Criddle, C. S. Fluorinated organics in the biosphere. Environ. Sci. Technol. 1997, 31, 2445−2454. (2) Schultz, M. M.; Barofsky, D. F.; Field, J. A. Fluorinated alkyl surfactants. Environ. Eng. Sci. 2003, 20, 487−501. (3) Giesy, J. P.; Kannan, K. Perfluorochemical surfactants in the environment. Environ. Sci. Technol. 2002, 36, 146A−152A. (4) Giesy, J. P.; Kannan, K. Global distribution of perfluorooctane sulfonate and related perfluorinated compounds in wildlife. Environ. Sci. Technol. 2001, 35, 1339−1342. (5) Hogue, C. PFOA called likely human carcinogen. Chem. Eng. News 2006, 83, 5. (6) Sanderson, H.; Boudreau, T. M.; Mabury, S. A.; Solomon, K. R. Impact of perfluorooctanoic acid on the structure of the zooplankton community in indoor microcosms. Aquat. Toxicol. 2003, 62, 227−234. (7) Hori, H.; Yamamoto, A.; Hayakawa, E.; Taniyasu, S.; Yamashita, N.; Kutsuna, S. Efficient decomposition of environmentally persistent perfluorocarboxylic acids by use of persulfate as a photochemical oxidant. Environ. Sci. Technol. 2005, 39, 2383−2388. (8) Kutsuna, S.; Hori, H. Rate constants for aqueous-phase reactions of SO4− with C2F5C(O)O− and C3F7C(O)O− at 298 K. Int. J. Chem. Kinet. 2007, 39, 276−288. (9) Hori, H.; Hayakawa, E.; Einaga, H.; Kutsuna, S.; Koike, K.; Ibusuki, T.; Kiatagawa, H.; Arakawa, R. Decomposition of environmentally persistent perfluorooctanoic acid in water by photochemical approaches. Environ. Sci. Technol. 2004, 38, 6118−6124. (10) Hori, H.; Yamamoto, A.; Koike, K.; Kutsuna, S.; Osaka, I.; Arakawa, R. Photochemical decomposition of environmentally persistent short-chain perfluorocarboxylic acids in water mediated by iron(II)/(III) redox reactions. Chemosphere 2007, 68, 572−578. (11) Wang, Y.; Zhang, P.; Pan, G.; Chen, H. Ferric ion mediated photochemical decomposition of perfluorooctanoic acid (PFOA) by 254 nm UV light. J. Hazard. Mater. 2008, 160, 181−186. (12) Park, H.; Vecitis, C. D.; Cheng, J.; Choi, W.; Mader, B. T.; Hoffmann, M. R. Reductive defluorination of aqueous perfluorinated alkyl surfactants: effects of ionic headgroup and chain length. J. Phys. Chem. A 2009, 113, 690−696. (13) Moriwaki, H.; Takagi, Y.; Tanaka, M.; Tsuruho, K.; Okitsu, K.; Maeda, Y. Sonochemical decomposition of perfluorooctane sulfonate and perfluorooctanoic acid. Environ. Sci. Technol. 2005, 39, 3388− 3392. (14) Cheng, J.; Vecitis, C. D.; Park, H.; Mader, B. T.; Hoffmann, M. R. Sonochemical degradation of perfluorooctane sulfonate (PFOS) and perfluorooctanoate (PFOA) in landfill groundwater: environmental matrix effects. Environ. Sci. Technol. 2008, 42, 8057−8063. (15) Vecitis, C. D.; Park, H.; Cheng, J.; Mader, B. T.; Hoffmann, M. R. Enhancement of perfluorooctanoate and perfluorooctanesulfonate activity at acoustic cavitation bubble interfaces. J. Phys. Chem. C 2008, 112, 16850−16857. (16) Vecitis, C. D.; Park, H.; Cheng, J.; Mader, B. T.; Hoffmann, M. R. Kinetics and mechanism of the sonolytic conversion of the aqueous perfluorinated surfactants, perfluorooctanoate (PFOA), and perfluor-

concentration was nearly 50 times higher than PFOA concentration added in the secondary effluent; the surface of In2O3 was mostly occupied by bicarbonate. As a result, little PFOA was adsorbed on In2O3, and its decomposition was accordingly inhibited. When the pH value of the secondary effluent was adjusted to 4, the adsorption of PFOA on In2O3 increased to some extent, which can be attributed to transformation of bicarbonate into carbonic acid at pH 4. However, PFOA decomposition was still significantly delayed, reflecting the impacts of organic matters in the secondary effluent. Considering that ozone addition can greatly accelerate the photocatalytic removal of organic matters,48,49 we investigated the effect of ozone addition on the PFOA decomposition in the wastewater. As shown in Figure 7, when the pH value of the secondary effluent was adjusted to 4 and ozone gas was simultaneously added (∼8 mg/L.h), PFOA decomposed almost as fast as that in the pure water. The above results indicate that PFOA decomposition in wastewater may be greatly inhibited by bicarbonate and organic matters; however, their impacts can be mostly avoided via pH adjustment and ozone addition. As compared with other existing technologies for PFOA decomposition,19 In2O3 photocatalysis works efficiently under mild conditions, i.e. room temperature, atmospheric pressure, and weak acidic condition. Moreover, its energy consumption is not high (7.6 kJ/μmol/L PFOA), and it has rather high reaction rate constant and short half-life time, i.e. 110 min for 100 μmol/L PFOA. More importantly, this study provides a new method to decompose PFOA and other PFCAs. The In2O3 material used in this study is a commercial product with low surface area and large particle size. As well-known, activity of a photocatalyst can be significantly increased through improvement of material synthesis, such as control of morphology and size,50,51 crystal facet engineering, and surface modification.44,51−54 Thus, development of new In2O3 nanomaterials with higher activity should be considered as well as its applicability to remove trace-level PFOA in wastewater.

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ooctane sulfonate (PFOS) into inorganic products. J. Phys. Chem. A 2008, 112, 4261−4270. (17) Cheng, J.; Vecitis, C. D.; Park, H.; Mader, B. T.; Hoffmann, M. R. Sonochemical degradation of perfluorooctane sulfonate (PFOS) and perfluorooctanoate (PFOA) in groundwater: kinetic effects of matrix inorganics. Environ. Sci. Technol. 2010, 44, 445−450. (18) Kutsuna, S.; Nagaoka, Y.; Takeuchi, K.; Hori, H. TiO2-induced heterogeneous photodegradation of a fluorotelomer alcohol in air. Environ. Sci. Technol. 2006, 40, 6824−6829. (19) Vecitis, C. D.; Park, H.; Cheng, J.; Mader, B. T.; Hoffmann, M. R. Treatment technologies for aqueous perfluorooctanesulfonate (PFOS) and perfluorooctanoate (PFOA). Front. Environ. Sci. Eng. China 2009, 3, 129−151. (20) Czili, H.; Horvath, A. Photodegradation of chloroacetic acids over bare and silver-deposited TiO2: Identification of species attacking model compounds, a mechanistic approach. Appl. Catal., B 2009, 89, 342−348. (21) Choi, W.; Hoffmann, M. R. Novel photocatalytic mechanisms for CHCl3, CHBr3, and CCl3CO2− degradation and the fate of photogenerated trihalomethyl radicals on TiO2. Environ. Sci. Technol. 1997, 31, 89−95. (22) Chemseddine, A.; Boehm, H. P. A study of the primary step in the photochemical degradation of acetic acid and chloroacetic acid on a TiO2 photocatalyst. J. Mol. Catal. 1990, 60, 295−311. (23) Panchangam, S. C.; Lin, A. Y.; Shaik, K. L.; Lin, C. F. Decomposition of perfluorocarboxylic acids (PFCAs) by heterogeneous photocatalysis in acidic aqueous medium. Chemosphere 2009, 77, 242−248. (24) Dillert, R.; Bahnemann, D.; Hidaka, H. Light-induced degradation of perfluorocarboxylic acids in the presence of titanium dioxide. Chemosphere 2007, 67, 785−792. (25) Chen, X.; Xue, H.; Li, Z. H.; Wu, L.; Wang, X. X.; Fu, X. Z. Ternary wide band gap p-block metal semiconductor ZnGa2O4 for photocatalytic benzene degradation. J. Phys. Chem. C 2008, 112, 20393−20397. (26) Yan, T. J.; Long, J. L.; Chen, Y. S.; Wang, X. X.; Li, D. Z.; Fu, X. Z. Indium hydroxide: A highly active and low deactivated catalyst for photoinduced oxidation of benzene. C. R. Chimie 2008, 11, 101−106. (27) Moody, C. A.; Field, J. A. Determination of perfluorocarboxylates in groundwater impacted by fire-fighting activity. Environ. Sci. Technol. 1999, 33, 2800−2806. (28) Lin, A. Y.; Panchangam, S. C.; Lo, C. The impact of semiconductor, electronics and optoelectronic industries on downstream perfluorinated chemical contamination in Taiwanese rivers. Environ. Pollut. 2009, 157, 1365−1372. (29) Rayne, S.; Forest, K. Perfluoroalkyl sulfonic and carboxylic acids: A critical review of physicochemical properties, levels and patterns in waters and wastewaters, and treatment methods. J. Environ. Sci. Health, Part A: Toxic/Hazard. Subst. Environ. Eng. 2009, 44, 1145−1199. (30) Park, J. W.; Seo, Y. A.; Kim, I.; Ha, C. S.; Aimi, K.; Ando, S. Investigating the crystalline structure of poly(vinylidene fluoride) (PVDF) in PVDF/silica binary and PVDF/poly(methyl methacrylate)/silica ternary hybrid composites using FTIR and solid-state 19F MAS NMR spectroscopy. Macromolecules 2004, 37, 429−436. (31) Cao, M. H.; Wang, B. B.; Yu, H. S.; Wang, L. L.; Yuan, S. H.; Chen, J. Photochemical decomposition of perfluorooctanoic acid in aqueous periodate with VUV and UV light irradiation. J. Hazard. Mater. 2010, 179, 1143−1146. (32) Hoffmann, M. R.; Martin, S. T.; Choi, W.; Bahnemann, D. W. Environmental applications of semiconductor photocatalysis. Chem. Rev. 1995, 95, 69−96. (33) Moser, J.; Punchihewa, S.; Infelta, P. P.; Gratzel, M. Surface complexation of colloidal semiconductors strongly enhances interfacial electron-transfer rates. Langmuir 1991, 7, 3012−3018. (34) Rotzinger, F. P.; Kesselman-Truttmann, J. M.; Hug, S. J.; Shklover, V.; Gratzel, M. Structure and vibrational spectrum of formate and acetate adsorbed from aqueous solution onto the TiO2 rutile (110) surface. J. Phys. Chem. B 2004, 108, 5004−5017.

(35) Franch, M. I.; Ayllon, J. A.; Peral, J.; Domenech, X. Enhanced photocatalytic degradation of maleic acid by Fe(III) adsorption onto the TiO2 surface. Catal. Today 2005, 101, 245−252. (36) Meulen, T.; Mattson, A.; Osterlund, L. A comparative study of the photocatalytic oxidation of propane on anatase, rutile, and mixedphase anatase-rutile TiO2 nanoparticles: Role of surface intermediates. J. Catal. 2007, 251, 131−144. (37) Infrared and Raman spectra of inorganic and coordination compounds, 4th, ed.; Nakamoto, K., Ed.;John Wiley & Sons Press: New York, 2009. (38) Deacon, G. B.; Phillips, R. J. Relationships between the carbonoxygen stretching frequencies of carboxylato complexes and the type of carboxylate coordination. Coord. Chem. Rev. 1980, 33, 227−250. (39) Buchanan, G. W.; Munteanu, E.; Dawson, B. A.; Hodgson, D. Concerning the origin of 19F-19F NMR COSY and NOESY connections in the spectra of perfluorooctanoic acid, RF-palmitic acid-F13 and diethyl perfluorosuberate. Magn. Reson. Chem. 2005, 43, 528−534. (40) Guide to fluorine NMR for organic chemists; Dolbier, W. R., Ed.; John Wiley & Sons Press: NJ, 2009. (41) Pawsey, S.; Reven, L. 19F fast magic-angle spinning NMR studies of perfluoroalkanoic acid self-assembled monolayers. Langmuir 2006, 22, 1055−1062. (42) Geppi, M.; Borsacchi, S.; Mollica, G.; Veracini, C. A. Applications of solid-state NMR to the study of organic/inorganic multicomponent materials. Appl. Spectrosc. Rev. 2009, 44, 1−89. (43) Yu, J. C.; Ho, W. K.; Yu, J. G.; Hark, S. K.; Iu, K. Effects of trifluoroacetic acid modification on the surface microstructures and photocatalytic activity of mesoporous TiO2 thin films. Langmuir 2003, 19, 3889−3896. (44) Yu, J. G.; Wang, W. G.; Cheng, B.; Su, B. L. Enhancement of photocatalytic activity of mesoporous TiO2 powders by hydrothermal surface fluorination treatment. J. Phys. Chem. C 2009, 113, 6743−6750. (45) Yang, S.; Sun, C.; Li, X.; Gong, Z.; Quan, X. Enhanced photocatalytic activity for titanium dioxide by co-modifying with silica and fluorine. J. Hazard. Mater. 2010, 175, 258−266. (46) Schroder, H. F.; Meesters, R. J. W. Stability of fluorinated surfactants in advanced oxidation processes - a follow up of degradation products using flow injection − mass spectrometry, liquid chromatography − mass spectrometry and liquid chromatography − multiple stage mass spectrometry. J. Chromatogr., A 2005, 1082, 110− 119. (47) Kosmulski, M. Pristine points of zero charge of gallium and indium oxides. J. Colloid Interface Sci. 2001, 238, 225−227. (48) Zhang, P.; Jian, L. Ozone-enhanced photocatalytic degradation of natural organic matter in water. Water Sci. Technol.: Water Supply 2006, 6, 53−61. (49) Li, L.; Zhang, P.; Zhu, W.; Han, W.; Zhang, Z. Comparison of O3-BAC, UV/O3-BAC and TiO2/UV/O3-BAC processes for removing organic pollutants in secondary effluents. J. Photochem. Photobiol., A 2005, 171, 145−151. (50) Chen, D.; Huang, F.; Cheng, Y.; Caruso, R. A. Mesoporous anatase TiO2 beads with high surface areas and controllable pore sizes: a superior candidate for high-performance dye-sensitized solar cells. Adv. Mater. 2009, 21, 2206−2210. (51) Chen, X.; Mao, S. S. Titanium dioxide nanomaterials: synthesis, properties, modifications, and applications. Chem. Rev. 2007, 107, 2891−2959. (52) Yang, H. G.; Sun, C. H.; Qiao, S. Z.; Zou, J.; Liu, G.; Smith, S. C.; Cheng, H. M.; Lu, G. Q. Anatase TiO2 single crystals with a large percentage of reactive facets. Nature 2008, 453, 638−642. (53) Liu, G.; Yu, J. C.; Lu, G. Q.; Cheng, H. Crystal facet engineering of semiconductor photocatalysts: motivations, advances and unique properties. Chem. Commun. 2011, 47, 6763−6783. (54) Fu, P.; Zhang, P.; Li, J. Photocatalytic degradation of low concentration formaldehyde and simultaneous elimination of ozone by-product using palladium modified TiO2 films under UV254+185nm irradiation. Appl. Catal., B 2011, 105, 220−228.

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