Photodegradation of Perfluorooctane Sulfonate by UV Irradiation in

National Institute for Environmental Studies, 16-2, Onogawa, Tsukuba, Ibaraki 305-8506, Japan, Kyoto University, Environment Preservation Center, ...
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Environ. Sci. Technol. 2007, 41, 5660-5665

Photodegradation of Perfluorooctane Sulfonate by UV Irradiation in Water and Alkaline 2-Propanol T A K A S H I Y A M A M O T O , * ,† Y U K I O N O M A , † SHIN-ICHI SAKAI,‡ AND YASUYUKI SHIBATA† National Institute for Environmental Studies, 16-2, Onogawa, Tsukuba, Ibaraki 305-8506, Japan, Kyoto University, Environment Preservation Center, Yoshida-Honmachi, Sakyo-ku, Kyoto 606-8501, Japan

Perfluorooctane sulfonate (PFOS) is the environmentally concerned compound because of its persistence and bioaccumulative properties. Since photodegradation of PFOS is not yet experimentally confirmed, photodegradation study of PFOS in water and alkaline 2-propanol solution was conducted. Aqueous and alkaline 2-propanol solution of PFOS (40 µM) was irradiated with a low-pressure mercury lamp (254 nm, 32 W) by internal irradiation for 10 d, and then PFOS, fluoride and sulfate ions, and the other degradation products were analyzed. Photodegradation of PFOS was confirmed in both media. PFOS was degraded by 8% after 1 day and by 68% after 10 days irradiation compared to the initial concentration in water. In alkaline 2-propanol, 76 and 92% of PFOS was degraded after 1 and 10 days irradiation, respectively. Photodegradation of PFOS in alkaline 2-propanol was much faster and effective than in water, as the photodegradation rate constants were 0.93 days-1 in alkaline 2-propanol and 0.13 days-1 in water, respectively. Formation of fluoride and sulfate was also confirmed by ion chromatography and X-ray diffraction analysis. From observation of the degradation products, two major degradation pathways of PFOS were considered: via C8HF17 and C8F17OH, respectively, resulting in short-chain fluorinated compounds such as C7HF15 and C7F15OH by stepwise removal of CF2. Formation of shortchain fluorocarbons such as CF4, C2F6, and C3F8 were also confirmed. This is the first study to confirm photodegradation of PFOS in water and alkaline 2-propanol.

Introduction Perfluorinated sulfonic and carboxylic acids are of special public concern as environmental contaminants, because they are globally distributed, persistent, and bioaccumulative for higher chain homologues. Perfluorooctane sulfonate (PFOS) is one of the main types of perfluorinated chemicals, and has been widely detected in humans, wildlife, and the environment (1-5). PFOS and related chemicals have been employed in industrial and commercial applications for approximately 50 years as additives in, for example, surface treatment; paper protection, and performance chemicals for clothing, carpets, * Corresponding author phone: +81-29-850-2547; fax: +81-29850-2840; e-mail: [email protected]. † National Institute for Environmental Studies. ‡ Kyoto University. 5660

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leather, food plates, and containers; fire fighting foams; coating additives; cleaner products; pesticides; and so on. By the end of 2000, about 90% of 3M’s production of perfluorooctanyl-based chemicals was voluntarily phased out, and by the beginning of 2003, production completely ceased (6, 7). Over the peak production year in 2000, the global production volume for surface treatment applications, paper protection, and performance chemical applications was estimated at approximately 2160, 1490, and 831 metric tons, respectively (8), and the global production by 3M prior to cessation was estimated at 13 670 metric tons (1985-2002) (7). Most solid wastes containing PFOS are either incinerated or disposed of in the landfills; however, complete destruction of PFOS and related substances is favored because of its low rate of decomposition and the fact that the exact route of decomposition is unknown at present. Development of decomposition techniques for the large amount of PFOS and related chemicals produced over the past half a century are therefore required. Stockpiles and wastewater containing fluorinated compounds also may need to be destroyed while taking care to minimize secondary pollution. Ellis et al. showed the possibility of haloacetic acid production during combustion of PFOS and related compounds through the thermolysis of fluorinated polymers (9); on the contrary, Yamada et al. reported that polyester/ cellulose fabric containing fluorotelomer-based acrylic polymer could be destroyed at a temperature of 1000 °C without formation of perfluorooctanoic acid (PFOA) (10). Regardless, even though the atmospheric source of fluorinated compounds is not clearly known at present, a very high temperature is certainly a necessity during the combustion of fluorinated products (11). Other destruction technologies have also been studied such as sonochemical decomposition and subcritical water oxidation (12, 13). Photolysis is another possibility; however, according to a report by 3M, no evidence of direct photolysis was observed over a period of 167 h in the presence of hydrogen peroxide, iron oxide, and humic material at environmental concentrations (14). Hidaka et al. also reported that PFOS was not photodecomposed when used for titania surface modification (15, 16). The lack of, or very slow decomposition of PFOS is, therefore, thought to be due to its very strong carbon-fluorine binding property. Nevertheless, Hori et al. managed to successfully decompose PFOA in water using UV with heteropolyacid photocatalyst and persulfate (17, 18), which suggests the possibility of decomposition by photolysis. In the present study, therefore, we examined the decomposition of PFOS by internal UV irradiation in water and in alkaline 2-propanol which has already been used to treat waste PCB (19).

Experimental Section Materials. Perfluorooctane sulfonic acid (>98%) was purchased from Tokyo Kasei Kogyo Co. (Tokyo, Japan). Trifluoromethane (CHF3), tetrafluoromethane (CF4), hexafluoroethane (C2F6), octafluoropropane (C3F8) were purchased from Takachiho Chemical Industrial Co. (Tokyo, Japan). Perfluorohexane (C6F14), perfluoroheptane (C7F16), 1H-perfluorooctane (C8HF17) and perfluorooctane (C8F18) were purchased from Sigma-Aldrich Co. (Mo, US). 2-Propanol, sodium hydroxide, hydrochloric acid, sodium acetate, sodium carbonate, and sodium bicarbonate were of reagent grade, and methanol, acetone, and acetonitrile were of pesticide analysis grade. Purified water was prepared using the Millipore Milli-Q Gradient water purification system (MA). 10.1021/es0706504 CCC: $37.00

 2007 American Chemical Society Published on Web 07/07/2007

24 h using a gastight syringe and analyzed by GC/MS. The reaction temperature was within a range of 36-46 °C during the experiments. A control experiment in the absence of light irradiation was also carried out. Photodegradation in Alkaline 2-Propanol. The photodegradation experiments in the alkaline 2-propanol solutions were performed using the same procedures as described above, except for the solvents. The reaction solution contained PFOS at 20 mg L-1 (40 µM) was prepared in alkaline 2-propanol with sodium hydroxide (2.72 g, 68 mmol). The reaction temperature was within a range of 38-50 °C during the experiments. During the photodegradation period, deposition of white precipitates was observed. After the reaction, these precipitates were separated and analyzed by X-ray diffractometry. A control experiment without light irradiation was also conducted.

FIGURE 1. Schematic view of the photoreaction system. For cleanup of PFOS in the alkaline 2-propanol, a Waters SepPak Vac tC18 cartridge was used. Standard stock solution for calibration of LC/MS analysis of PFOS was prepared by dissolving of 10 mg of perfluorooctane sulfonic acid to 10 mL of methanol. Calibration standard solutions were prepared by stepwise dilution of standard stock solution. Standard solution for calibration of ion chromatography was purchased from Wako Pure Chemical Industries Inc. (Osaka, Japan). Apparatus. A photoreactor (UVL-32LB-1000Q; Riko Kagaku, Chiba, Japan) equipped with a low-pressure mercury lamp (UVL-32LB; 254 nm, 32W) was used. Photodegradation experiments were performed by internal irradiation, whereby a UV lamp was sunk into the reaction solution and the solution was irradiated internally. UV intensity on the surface of the lamp sheath was 3.73 mW cm-2 as recorded with a Toray Techno UV-500 UV meter (Shiga, Japan) equipped with a T-254 sensor. For quantitative analysis of PFOS, an Agilent 1100 LC/MSD operated in negative electrospray ionization (ESI) mode was used (see later for precise description of the LC/MS conditions). For qualitative analysis of photodegradation products in the gas phase, an Agilent 6890 GC/5973 MSD was used (see later for precise description of the GC/MS conditions). Fluoride and sulfate ions in the aqueous reaction solution were quantified with a Dionex DX-320 ion chromatograph (CA). Qualitative analysis of precipitates separated from the alkaline 2-propanol reaction solution was performed with a Rigaku RINT-Ultima X-ray diffractometer (Tokyo, Japan) using the Cu-KR line. Photodegradation in Water. A total of 750 mL of reaction solution containing PFOS (20 mg L-1, 40µM) was transferred to a photoreactor and purged for several min with N2 to remove O2. The reaction solution was then stirred and irradiated with a UV lamp placed in the center of the photoreactor for 10 days. All experiments were carried out at room temperature. A schematic view of the photoreactor is shown in Figure 1. The reaction solution was sampled every 24 h and analyzed by LC/MS and ion chromatography. Headspace gas of the photoreactor was also sampled every

LC/MS Analysis. PFOS in the reaction solution was quantified by LC/MS as previously described (20). Briefly, the aqueous reaction solution (0.5 mL) was diluted to 100 mL with methanol then analyzed by LC/MS. The alkaline 2-propanol reaction solution was neutralized with 0.05 M hydrochloric acid, diluted with purified water to 10 mL, then loaded onto a SepPak Vac tC18 cartridge column. The column was washed with 50 mL of 40% methanol/water then recovered with 50 mL of methanol. The methanol fraction was diluted to 100 mL and analyzed by LC/MS. Measurement conditions were as follows. Column: Zorbax Eclipse XDB C8 (4.6 mm i.d. × 150 mm, 5 µm); mobile phase: 10 mM sodium acetate/acetonitrile (55: 45); flow rate: 0.2 mL min-1; column temperature: 25 °C; injection volume: 5 µL; ionization mode: negative electrospray ionization (ESI); nebulizer gas: N2 (50 psi); drying gas: N2 (10 l min-1); fragmentor voltage: 220 V; capillary voltage: 4000 V; ions for single ion monitoring (SIM): m/z 499 (M-H)-. PFOS was quantified by external calibration. A range of calibration curve was 1-1000 ng mL-1 (r2 ) 0.9994). Detection limit of PFOS was 0.87 ng mL-1. When the solution was injected directly by a syringe pump attached to the MS system, the flow rate and scan time were 10 µL min-1 and 1.0 s, respectively. Mass spectra were obtained from the average of 30 scans with a scan range of m/z 10-600. GC/MS Analysis. Headspace gas was directly injected and analyzed by GC/MS. GC/MS conditions were as follows. Capillary column: Chrompack PoraPLOT Q (0.32 mm i.d. × 25 m, df)10 µm); column temperature: 150 °C (30 min); carrier gas: 1.6 mL min-1; injection mode: split (50:1); injection volume: 100 µL; injector temperature: 220 °C; MS interface temperature: 230 °C; ionization: EI (70 eV); scan range: m/z 15 - 550; SIM ions: m/z 69. Ion Chromatography Analysis. Fluoride and sulfate ions in the aqueous reaction solution were quantified by ion chromatography. Conditions were as follows. Column: Dionex IonPac AS12A (4 mm i.d. × 200 mm), mobile phase: 2.7 mM Na2CO3/0.3 mM NaHCO3, flow rate: 1.0 mL min-1; suppressor: ASRA (in recycle mode, 50 mA); detection: conductivity. Fluoride and sulfate ions were quantified by external calibration. A range of calibration curve was 0.2-20 µg mL-1 for fluoride (r2 ) 0.9923) and 1-100 µg mL-1 for sulfate (r2 ) 0.9932). X-ray Diffraction Analysis. Precipitates suspended in the alkaline 2-propanol solution after 10 days of UV irradiation were separated by centrifugation (2000 rpm, 10 min), washed with 2-propanol and acetone, and dried. Yield of precipitates was 0.12 g. Precipitates were analyzed using a X-ray diffraction meter. The conditions were as follows. Radiation: Cu-KR line; tube voltage: 40 kV; tube current: 26 mA; scan speed: 1° min-1; scan range: 2-90°. VOL. 41, NO. 16, 2007 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 2. Chromatogram change of PFOS isomer during photodegradation in water. Peak 4 was assigned as linear isomer, and peaks 1-3 were also assigned as branched ones.

FIGURE 3. Concentrations of PFOS, fluoride ion, and sulfate ion during photodegradation in water. UV energy was calculated using the following equation, UV energy(kJ/g) ) UV intensity per cm2(W cm-2) × effective surface area of UV lamp(cm2) × irradiation time(s) ÷ initial amount of PFOS(g).

Results and Discussion Photodegradation in Water. Chromatogram change of PFOS isomers during the photodegradation experiment in water is presented in Figure 2. Technical PFOS contains linear and branched isomers (21). As shown in Figure 2, four peaks were detected in our LC/MS analysis, the largest peak eluted at the last (peak 4) was assigned as a linear PFOS isomer and the others (peak 1-3) were assigned as branched isomers. We will mainly discuss the linear isomer here. PFOS concentration decreased by 8% after 1 day and 68% after 10 days compared to the initial concentration. It was confirmed that this decrease in PFOS concentration was not due to adsorption or vaporization, since the concentration did not change during the control experiment in the absence of light irradiation. Concentrations of fluoride and sulfate ions increased during the photodegradation experiment (Figure 3), suggesting that C-F and C-S bonds in the PFOS molecule dissociated by irradiation resulting in formation of related ions. From these results, it was confirmed that PFOS could be photodecomposed by UV irradiation in water. When one mole of PFOS is completely decomposed, 17 mol of fluoride and one mole of sulfate are stoichiometrically formed. After 10 days irradiation, the decrease of PFOS concentration was 27.2 µM and the increase of sulfate ions was 24.6 µM. Since the molar ratio of the formed sulfate ions 5662

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FIGURE 4. Concentrations of PFOS during photodegradation in alkaline 2-propanol. to decomposed PFOS was 0.91 (nearly 1), the presence of sulfur-containing products other than sulfate ions is thought to be limited. As for formation fluoride ions, the increase after 10 days irradiation was 326 µM. The increase of fluoride ions from the decrease of PFOS was stoichiometrically 462 µM ()27.2 µM × 17). The molar ratio of the observed increase to the expected one was 0.71, suggesting that 71% of fluorine contained in PFOS was converted to fluoride ions. The remainder is thought to have formed short-chain organofluorine compounds such as fluorocarbons, alcohols, and carboxylates. As for photodegradation of perfluorocompounds in aqueous media, Hori et al. (17, 18) reported photodegradation of PFOA with and without a photocatalyst such as heteropolyacid (H3PW12O40‚6H2O) or persulfate (S2O82-). In their study, PFOA solution was irradiated with a xenon-mercury lamp (220 W) placed outside the reaction vessel. On one hand, they reported that 95.9% of initial PFOA (1.35 mM) was decomposed by direct photolysis after 72 h irradiation. On the other hand, Chen and Zhang reported that only 4.8% of initial PFOA (60 µM) were decomposed by direct photolysis at 254 nm light (22). They used a 23 W low-pressure mercury lamp as a light source which was placed in the center of a reactor. In our preliminary experiments, photodegradation of PFOA was approximately 1.7 times faster than ones of PFOS. Therefore, with enough irradiation time and by proper irradiation method, PFOA can be decomposed by UV irradiation.

FIGURE 5. X-ray diffraction patterns of the precipitation separated from alkaline 2-propanol. Peaks of sodium fluoride are indicated by circles.

FIGURE 6. Photodegradation rates of PFOS isomers in alkaline 2-propanol. Peaks are numbered as elution order in LC/MS analysis (see Figure 2). As for photodegradation of PFOS, to the best of our knowledge, no previous reports have confirmed the photodegradation of PFOS in water. Hori et al. reported that PFOS (0.37 mM) was not decomposed by 12 h irradiation, whereas 36.8% of the initial concentration of PFOA was decomposed under the same conditions (13). Schro¨der and Meesters treated PFOS and other fluorinated surfactant by advanced oxidation processes (AOP), but PFOS was not decomposed (23). In these works, PFOS solution was irradiated by external irradiation, whereas PFOS solution was irradiated by internal irradiation in our study. Since internal irradiation should be more effective than external irradiation, decomposition of PFOS was confirmed in our study. Photodegradation in Alkaline 2-Propanol. During the photodegradation experiment in alkaline 2-propanol, the PFOS concentration decreased dramatically after 1 day, and then gradually over the next 9 days (Figure 4). Compared to the initial rate, the final concentration of PFOS decreased by 76% after 1 day and 92% after 10 days. Since PFOS concentration did not change in the absence of light irradiation, photodegradation of PFOS in alkaline 2-propanol was confirmed. X-ray diffraction analysis of the precipitates formed during the photodegradation experiment revealed diffraction peaks of sodium fluoride (NaF) in addition to peaks of sodium carbonate (Na2CO3) (Figure 5). Therefore, during photodegradation in alkaline 2-propanol, C-F bonds in the PFOS molecule are thought to be cleaved, forming sodium fluoride. We also found four peaks in the chromatograms of the reaction solution that were the same as degradation in water (Figure 2). The photodegradation rate constants of these four peaks, calculated assuming that the photodegradation reaction followed pseudo-first-order kinetics, were 1.0 (peak 1), 2.5 (peak 2), 1.8 (peak 3), and 0.93 days-1 (peak 4, linear isomer), respectively (Figure 6). Photodegradation of some

branched isomers (peaks 2 and 3) was faster than ones of linear isomer (peak 4). We will also discuss linear isomers in photodegradation in alkaline 2-propanol. Since the photodegradation rate of PFOS in alkaline 2-propanol was about 9.5 times higher than that in water, UV irradiation in alkaline 2-propanol solution is thought to be a possibility for the treatment of wastewater and stockpiles containing PFOS. This photodecomposition technique has already been used to treat waste PCB (19). Photodechlorination of PCBs progresses through a radical chain reaction mediated by 2-hydroxypropyl radicals initially formed by UV irradiation (24). Because photodegradation of PFOS in alkaline 2-propanol was much faster and effective than those in water, it seems that the same radical defluorination mechanism might work. The estimated photodegradation rate constants were 0.13 days-1 in water and 0.93 days-1 in alkaline 2-propanol, respectively, and irradiation energy values to reduce 50% of initial amount of PFOS were about 20 000 kJ g-1 in water and 2640 kJ g-1 in alkaline 2-propanol, respectively. As for degradation techniques of PFOS, Hori et al. reported the subcritical water oxidation method with iron (13); PFOS was degraded at 350 °C and 23.3 MPa over 6 h. Moriwaki et al. also reported sonochemical decomposition of PFOS and PFOA in water (12). They used ultrasonic apparatus operating at 200 W, resulting in the degradation of 60 and 85% PFOS and PFOA over 60 min irradiation, respectively. UV irradiation in alkaline 2-propanol is thought to be one of the degradation techniques for PFOS stockpile, though degradation using this technique is milder than the other two methods. Photodegradation Products. Observation of the photodegradation products formed from standard aqueous PFOS solution before and after 1 day UV irradiation were performed by negative ESI/MS. Figure 7a and b show the mass spectra and assigned ionic compounds of standard and photodegraded solution. The peaks shown in Figure 7 (both 7a and b) are not thought to be fragment ions at the interface of the ESI/MS, because no fragmentation of the perfluorocompounds was observed under the same conditions in our preliminary experiments. Only PFOS peaks were observed and the ions shown in Figure 7b were not observed in the mass spectra of the PFOS standard solution (Figure 7a). The photodegradation solution gave peaks at m/z 435, 419, and 413, and were assigned to [C8F17O]-, [C8F17]-, and [C7F15COO]-, respectively. Their compositional formulas were C8F17OH, C8HF17, and C7F15COOH because a proton was lost during ionization by the ESI/MS. The peaks at m/z 369, 319, 269, 219, and 169 with dissociation of CF2 from [C8F17]- were assigned to [CnF2n+1]- (n ) 3-7). In the same way, a series of peaks at m/z 385, 335, 285, 235, and 185 resulting from dissociation of CF2 from [C8F17O]- could be assigned to VOL. 41, NO. 16, 2007 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 7. Mass spectra with negative ESI and assigned ionic compounds of PFOS standard and photodecomposition products. (a) Aqueous PFOS standard solution before UV irradiation. (b) Aqueous solution of photodegradation products in water after 1 day UV irradiation. The intensity of the peaks at m/z 80-490 were magnified 10×.

FIGURE 8. SIM chromatograms (m/z 69) of headspace gas samples in photodegradation of PFOS. [CnF2n+1O]-, and a series of peaks at m/z m/z 363, 313, 263, 213, 163 resulting from dissociation of CF2 from [C7F15COO]also could be assigned to [Cn-1F2n-1COO]- (n ) 3-7). C8HF17 was thought to have been formed by the cleavage of C-S bonds in the PFOS and addition of a proton. In the same way, C8F17OH was produced by the cleavage of C-S bonds in the PFOS and addition of OH. According to Hori et al. (17), C7F15 radicals, produced by the cleavage of C-C bonds between C7F15 and COOH, form C7F15OH. This alcohol then undergoes HF elimination to form C6F13COF, which undergoes hydrolysis giving C6F13COOH with one less CF2 unit. A similar mechanism may, therefore, have occurred during the formation of short-chain perfluorocarboxylic acids by stepwise removal of CF2 from C8F17OH. This alcohol, however, is also thought to have undergone 2F elimination to form C7F15CHO, which showed an ion peak at m/z 397 in the mass spectra, but not C7F15COF. Although a series of peaks at m/z 381, 331, and 281 were assigned to [CnF2n-1]- (n ) 6-8), we 5664

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were unable to determine how these ions were formed. Taken together, the above results suggest two major degradation pathways of PFOS, via C8HF17 and C8F17OH, respectively, resulting in the production of short-chain perfluorocompounds by stepwise removal of CF2. Moriwaki et al. (12) found peaks at m/z 449, 413, and 363, which were assigned to [C7F15SO3]-, [C7F15COO]- (PFOA), and [C6F13COO]-, respectively, in the sonochemical degradation solution of PFOS analyzed by negative ESI/MS. They suggested that PFOA was formed by sonochemical oxidation after cleavage of C-S bond in PFOS. They also found the peaks appeared at m/z 113, 163, 213, 263, and 313 which were assigned to [CF3(CF2)nCOO]- (n ) 0-4). These shortchain perfluorocarboxylic acids were also formed by repetition of dissociation of COO- group and oxidation. Since the peaks at m/z 449, 413, and 363 were also found in the photodegradation solution as same as in the sonochemical degradation solution, the photodegradation pathways of

PFOS in the early stage may be similar to the sonochemical degradation pathways. However, because of the lack of the peaks assigned to short-chain perfluorocarboxylic acids in the photodegradation solution, further photodegradation pathways may be different from the pathways of sonochemical degradation. Possible photodegradation pathways of PFOS in the early stage were estimated as follows:

(6) (7)

(8)

(9) (10) (11) (12)

When the headspace gas from the photoreactor was analyzed, degradation products such as short-chain fluorocarbons were thought to be in the gas phase. In headspace gas analysis by GC/MS operating in the full scan mode, no fluorocompounds were detected. It was difficult to detect the molecular ions of the fluorocompounds because fragmentation occurred easily due to electron-impact ionization. Therefore, typical mass fragments generated from the organofluorine compounds were monitored in the SIM mode. SIM chromatograms (m/z 69) of headspace gas samples obtained from the photodegradation experiment in water are presented in Figure 8. This fragment ion was derived from CF3, which was a typical fragment ion from perfluorocompounds. The peak appeared at 14 min in the chromatogram of the 1 day irradiated sample, and it was assigned as C7F16. The intensity of this peak decreased with increasing reaction period, and several peaks appeared in the retention time between 3 and 6 min. Comparison of the retention times of fluorocarbon standards allowed tentative assignment of the peaks at 2.8, 3.0, and 3.4 min to CF4, C2F6, and C3F8, respectively. Though these peaks were not identified by mass spectra, it is likely that short-chain fluorocarbons were formed by photodegradation. As shown in Figure 8, since the intensities of these peaks increased with reaction time, these compounds are thought to be formed from PFOS photodegradation. Since these short-chain fluorocarbons are classified as greenhouse gases, it is necessary to avoid their emission during PFOS photodegradation. Further study is needed to determine these decomposition products and to clarify the degradation pathway as well as the mass balance.

(13)

(14)

(15)

(16)

(17)

(18)

(19) (20)

(21)

Literature Cited (1) Olsen, G. W.; Church, T. R.; Larson, E. B.; Belle, G.; Lundberg, J. K.; Hansen, K. J.; Burris, J. M.; Mandel, J. H.; Zobel, L. R. Serum concentrations of perfluorooctanesulfonate and other fluorochemicals in an elderly population from Seattle, Washington. Chemosphere 2004, 54, 1599-1611. (2) Giesy, J. P.; Kannan, K. Global distribution of perfluorooctane sulfonate in wildlife. Environ. Sci. Technol. 2001, 35, 13391342. (3) Kannan, K.; Koistinen, J.; Beckmen, K.; Evans, T.; Gorzelany, J. F.; Hansen, K. J.; Jones, P. D.; Helle, E.; Nyman, M.; Giesy, J. P. Accumulation of Perfluorooctane sulfonate in marine mammals. Environ. Sci. Technol. 2001, 35, 1593-1598. (4) Moody, C. A.; Field, J. A. Determination of perfluorocarboxylates in ground water impacted by fire-fighting activity. Environ. Sci. Technol. 1999, 33, 2800-2806. (5) Martin, J. W.; Muir, D. C. G.; Moody, C. A.; Ellis, D. A.; Kwan, W. C.; Solomon, K. R.; Mabury, S. A. Collection of airborne

(22) (23)

(24)

fluorinated organics and analysis by gas chromatography/ chemical ionization mass spectrometry. Anal. Chem. 2002, 74, 584-590. Renner, R. Grown concern over perfluorinated chemicals. Environ. Sci. Technol. 2001, 35, 154A. UNEP. Draft risk profile: perfluorooctane sulfonate (PFOS). UNEP/POPS/POPRC.2/11, 2006, 9-10. http://www.unon.org/ confss/doc/unep/pops/POPRC_02/POPRC.2_11/ K0652253.pdf. OECD. Co-operation on Existing ChemicalssHazard Assessment of Perfluorooctane Sulfonate (PFOS) and its Salts; U.S. Environmental Protection Agency docket AR-226-1140; U.S. Environmental Protection Agency: Washington, DC 2002. Ellis, D. A.; Mabury, S. A.; Martin, J. W.; Muir, D. C. G. Thermolysis of fluoropolymers as a potential source of halogenated organic acids in the environment. Nature 2001, 412, 321-324. Yamada, T.; Taylor, P. H.; Buck, R. C.; Kaiser, M. A.; Giraud, R. J. Thermal degradation of fluorotelomer treated articles and related materials. Chemosphere 2005, 61, 974-984. Shultz, M. M.; Barofsky, D. F.; Field, J. A. Fluorinated alkyl surfactants. Environ. Eng. Sci. 2003, 20, 487-501. Moriwaki, H.; Takagi, Y.; Tanaka, M.; Tsuruho, K.; Okitsu, K.; Maeda, Y. Sonochemical decomposition of perfluorooctane sulfonate and perfluorooctanoic acid. Environ. Sci. Technol. 2006, 39, 3388-3392. Hori, H.; Nagaoka, Y; Yamamoto, A.; Sano, T.; Yamashita, N.; Taniyasu, S.; Kutsuna, S. Efficient decomposition of environmentally persistent perfluorooctanesulfonate and related fluorochemicals using zerovalent iron in subcritical water. Environ. Sci. Technol. 2006, 1049-1054. 3M Company. Environmental and Health Assessment of Perfluorooctane Sulfonic Acid and its Salts, St. Paul, MN; U.S. Environmental Protection Agency docket AR-226-1486; U.S. Environmental Protection Agency: Washington, DC, 2003. Hidaka, H.; Jou, H.; Nohara, K.; Zhao, J. Photocatalystic degradation of the hydrophobic pesticide permethrin in fluoro surfactant/TiO2 aqueous dispersions. Chemosphere 1992, 25, 1589-1597. Yuan, Q.; Raxikrishna, R.; Valsaraj, K. T. Reusable adsorbents for dilute solution separation. 5. Photodegradation of organic compounds on surfactant-modified titania. Sep. Purif. Technol. 2001, 24, 309-318. Hori, H.; Hayakawa, E.; Einaga, H.; Kutsuna, S.; Koike, K.; Ibusuki, T.; Kitagawa, H.; Arakawa, R. Decomposition of environmentally persistent perfluorooctanoic acid in water by photochemical approaches. Environ. Sci. Technol. 2004, 38, 6118-6124. 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, 23832388. Noma, Y.; Ohno, M.; Sakai, S. Pathways for the degradation of PCBs by palladium-catalyzed dechlorination. Fresenius Environ. Bull. 2003, 12, 302-308. Saito, N.; Harada, K.; Inoue, N.; Sasaki, K.; Yoshinaga, T.; Koizumi, A. Perfluorooctanoate and perfluorooctane sulfonate concentrations in surface water in Japan. J. Occup. Health 2004, 46, 49-59. Langlois, I.; Oehme, M. Structural identification of isomers present in technical perfluorooctane sulfonate by tandem mass spectrometry. Rapid Commun. Mass Spectrom. 2006, 20, 844850. Chen, J.; Zhang, P. Photodegradation of perfluorooctanoic acid in water under irradiation of 254 nm and 185 nm light by use of persulfate. Water Sci. Technol. 2006, 54 (11-12), 317-325. Schro¨der, H. Fr.; 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. Nishiwaki, T.; Ninomiya, A.; Yamanaka, S.; Anda, K. Dechlorination of polychlorinated biphenyls by UV-irradiation (in Japanese). J. Chem. Soc. Jpn. 1973, 12, 2326-2331.

Received for review March 15, 2007. Revised manuscript received May 31, 2007. Accepted June 7, 2007. ES0706504

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