Indirect Photolysis of Perfluorochemicals: Hydroxyl Radical-Initiated

Apr 20, 2009 - However, indirect photolysis may take place instead. ..... Portions of this research were funded by the Santa Clara Valley Water Distri...
21 downloads 0 Views 429KB Size
Download PDF
Environ. Sci. Technol. 2009, 43, 3662–3668

Indirect Photolysis of Perfluorochemicals: Hydroxyl Radical-Initiated Oxidation of N-Ethyl Perfluorooctane Sulfonamido Acetate (N-EtFOSAA) and Other Perfluoroalkanesulfonamides M E G A N H . P L U M L E E , †,§ KRISTOPHER MCNEILL,‡ AND M A R T I N R E I N H A R D * ,† Department of Civil and Environmental Engineering, Yang and Yamazaki Environment and Energy Building, 473 Via Ortega, Stanford University, Stanford, California 94305-4020, and Department of Chemistry, 207 Pleasant Street SE, University of Minnesota, Minneapolis, Minnesota 55455

Received December 2, 2008. Revised manuscript received March 16, 2009. Accepted March 18, 2009.

Selected perfluorinated surfactants were irradiated in aqueous hydrogen peroxide solutions using artificial sunlight to study transformation under aquatic environmental conditions. Indirect photolysis mediated by hydroxyl radical was observed for N-ethyl perfluorooctane sulfonamidoethanol (N-EtFOSE), N-ethylperfluorooctanesulfonamidoacetate(N-EtFOSAA),N-ethyl perfluorooctane sulfonamide (N-EtFOSA), and perfluorooctane sulfonamide acetate (FOSAA). An upper limit for the bimolecular reaction rate constant for reaction of •OH and N-EtFOSAA was determined to be (1.7 ( 0.7) × 109 M-1s-1. A proposed reaction pathway for degradation of the parent perfluorochemical, N-EtFOSE, to the other perfluoroalkanesulfonamides and perfluorooctanoate (PFOA) was developed and includes oxidation and N-dealkylation steps. As they did not undergo additional degradation, perfluorooctane sulfonamide (FOSA) and PFOA were the final degradation products of hydroxyl radical-initiated oxidation. UV-visible absorption spectra for the perfluorochemicals, showing absorbance in the UV region below the range of natural sunlight, are also reported. In sunlit environments, indirect photolysis of perfluorochemicals is likely to be important in the determination of their environmental fate given the slow rates expected for biotransformation and weak sorption. Photolytic conversion of perfluorochemicals into refractory perfluorinated acids, mainly PFOA, could mean that a significant fraction of these compounds will accumulate in the world’s oceans.

Introduction Perfluorochemicals (PFCs) are persistent environmental contaminants used in a variety of products including fluoropolymers, liquid repellents for paper and textiles, and * Corresponding author phone: (650) 723-0208; fax: (650) 7237058; e-mail: [email protected]. † Stanford University. § Present address: Exponent, 149 Commonwealth Drive, Menlo Park, California 94025, USA. ‡ University of Minnesota. 3662

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 10, 2009

industrial additives (1). They are found in water, air, and sediment (2-6) as well as in wildlife and humans (7-9). The carbon-fluorine bond is extremely strong, and thus degradation of the perfluorinated carbon-chains is generally not observed. Transformation of PFCs may occur by modification of the nonfluorinated carbon structure, as has been observed in biological studies (10, 11) and in atmospheric reactions (5, 12, 13). Biotransformation of PFCs is slow (10, 14) and sorption is weak: Rhoads et al. (10) observed half-lives ranging 0.71 to 9.2 days for biodegradation in activated sludge from a wastewater treatment plant. Higgins and Luthy (15) report log Koc values of 3.11 ( 0.16 and 3.23 ( 0.18 for N-methyl perfluorooctane sulfonamido acetate (N-MeFOSAA) and N-ethyl perfluorooctane sulfonamido acetate (N-EtFOSAA), respectively, which indicate that PFCs sorb relatively weakly to sediments. Volatilization is important for those PFCs without acidic functional groups, such as N-ethyl perfluorooctane sulfonamidoethanol (N-EtFOSE) (10). The potential for aquatic photolysis of PFCs due to sunlight irradiation has received limited attention. PFCs do not absorb within the spectrum of natural sunlight (λ > 290 nm) (16); therefore, direct photolysis of PFCs is not expected under typical environmental conditions. However, indirect photolysis may take place instead. Indirect photolysis of a target compound occurs via reaction with photochemically produced reactive intermediates (PPRIs). One important PPRI is the hydroxyl radical (•OH), which is found in freshwaters at steady-state concentrations ([•OH ]SS) ranging from 10-18 to 10-14 M (17, 18), in atmospheric water droplets (19), and in radical oxidation processes (advanced oxidation) for wastewater treatment (20, 21). Hydroxyl radicals typically react via hydrogen atom abstraction or addition to double bonds at rates near the diffusion-controlled limit (ca. 1010 M-1s-1) (21, 22). PFC photolysis research to date has focused on UV irradiation and/or advanced oxidation for wastewater treatment. Hori et al. (16) observed direct photolysis of PFOA by light absorption in the UV-region which was enhanced by a heteropolyacid photocatalyst; other studies have reported the degradation of PFCs using persulfate-induced indirect photolysis (23, 24). Complete decomposition to F- and CO2 was typically observed. More relevant to environmental conditions, Gauthier and Mabury (25) reported the aqueous •OH initiated oxidation of 8:2 fluorotelomer alcohol (8:2 FtOH) in irradiated hydrogen peroxide (H2O2) solutions (used to generate •OH), as well as in synthetic and natural lake water, leading to PFC products. In industry studies, neither direct nor indirect (•OH initiated oxidation using H2O2 irradiation) photolysis of PFOA or PFOS was observed under simulated environmental conditions (26, 27). Indirect photolysis (•OH initiated oxidation using H2O2 irradiation) of N-EtFOSE was observed forming N-ethyl perfluorooctane sulfonamide (N-EtFOSA), PFOA, and perfluorooctane sulfonamide (FOSA) (direct photolysis of N-EtFOSE was not observed) (28). The objectives of the present study were to investigate the potential for aqueous indirect photolysis of perfluorooctanesulfonamides, including N-EtFOSE and N-EtFOSAA, and to determine the reaction pathways and products. N-EtFOSAA has been detected in wastewater and wastewaterimpacted surface waters (3, 4, 29), and N-EtFOSE is also expected to occur in wastewater (10). Although some PFC manufacturers have transitioned from C8 to C4 perfluorinated compounds due to the greater extent of bioaccumulation observed for C8 compounds (30), the perfluorooctane-based chemicals that have been released are persistent and 10.1021/es803411w CCC: $40.75

 2009 American Chemical Society

Published on Web 04/20/2009

FIGURE 1. Absorbance spectra for selected PFCs in methanol. Irradiance of the photosimulator is shown on the secondary y-axis. ubiquitous. Because the perfluoroalkyl group is refractory, the photochemical pathways elucidated for the C8 compounds are expected to be analogous to those of the C4 analogues and degradation products.

Experimental Section Materials and Methods. Using large-volume injection, PFC samples were analyzed by liquid chromatography-tandem mass spectrometry (6, 10, 29) (see the Supporting Information (SI)). Irradiations were performed using an Atlas Suntest CPS+ photosimulator (Chicago, IL) equipped with a 1.1 kW xenon arc lamp at an intensity of 765 W/m2. The lamp was fitted with glass filters to block wavelengths below 290 nm to simulate natural sunlight (passing wavelength, 290 nm < λ < ∼800 nm; manufacturer-specified irradiance indicated in Figure 1). Irradiation Experiments. PFCs (5-15 µg/L initial concentration of N-EtFOSE, N-EtFOSAA, N-MeFOSAA, N-EtFOSA, FOSAA, FOSA, or PFOA) were irradiated individually in solutions of 10 mM H2O2 (Fisher Scientific; Pittsburgh, PA) in identical capped 20 mL quartz test tubes (Quartz Scientific, Fairport, OH) for intervals ranging from 1 to 6 days. Additional tubes containing the target PFC and H2O2 were covered with aluminum foil and irradiated simultaneously to serve as dark controls. Because the critical micelle concentrations (CMCs) of several of these PFCs are unknown, the experimental concentrations were chosen to be well below the CMC of PFOA (1-4 g/L) (31). H2O2 was used to generate hydroxyl radicals according to the reaction: hv

H2O2 98 2·OH The relatively high (10 mM) initial concentration of H2O2 (and consequently, hydroxyl radical) was required in order to observe significant decay of target PFCs over the experimental time period. Therefore, the decay rates observed were not representative of environmental conditions. The direct photolysis of aqueous 4-nitroanisole (PNA) was also monitored in a separate tube for quality control purposes as a measure of the consistency of lamp output. Aqueous PFC solutions for the irradiation experiments were prepared by spiking a small volume of a methanolbased PFC stock and allowing the methanol to evaporate before addition of Milli-Q water (Millipore, Billerica, MA) or HPLC-grade water. The solution was poured into the quartz tubes, which were completely filled to avoid headspace and

placed horizontally in a constant temperature (20 ( 2 °C) water bath 25 cm directly below the photosimulator lamp. Because PFCs are surface-active and therefore may accumulate at the air-water interface or on container surfaces in aqueous samples, special care was taken in the experimental design. First, controls were necessary not only to detect any unexpected losses of the starting compound, but also to confirm that a consistent initial concentration was achievable in the tubes following transfer from the aqueous stock solution. Second, rather than subsampling, an entire tube was sacrificed at each time point. Samples were taken by transferring the contents into a polypropylene microcentrifuge tube (E & K Scientific; Santa Clara, CA) containing methanol such that the final solution was 60/40 aqueous sample:methanol. Dilution with methanol serves to quench · OH and solvate the PFCs, allowing subsampling. Duplicate aliquots of the sample-methanol mixture were taken for analysis. Kinetic and Product Study. Where analytical standards were available, products formed were monitored and each product was irradiated individually to elucidate the reaction pathway. Products monitored included the perfluoroalkanesulfonamides from the present study, perfluorocarboxylates (C8 to C6), perfluorinated sulfonates (C8 to C6), and perfluorooctane sulfinate (PFOSI) (10), with instrumental detection limits typically between 0.002 and 0.05 µg/L. Fluorine mass balances were evaluated from the concentrations of measured perfluorochemicals. The observed pseudo first order decay rate of the starting compound was determined as the slope of ln (C0/C) vs time, with the standard error determined from linear regression analysis. As noted above, the laboratory decay rates measured are useful for relative purposes only, as the experimental concentration of •OH was not representative of environmental conditions. Hydroxyl Radical (•OH) Reaction. An upper limit for the bimolecular reaction rate constant for the reaction of •OH and N-EtFOSAA was determined using a variation of a method described by Boreen et al. (32). The degradation of acetophenone (a reference compound) and N-EtFOSAA (the substrate) were monitored (n ) 4 experiments) during aqueous irradiation with H2O2. Acetophenone (or its degradation products) appeared to interfere with the degradation of N-EtFOSAA; therefore, the two compounds were irradiated in separate solutions using identical experimental conditions. The initial (and excess) concentration of H2O2 is not expected to have decreased significantly (t1/2 > 3 days) over the course of the short experiment and therefore provided a constant VOL. 43, NO. 10, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

3663

FIGURE 2. Proposed pathway for the aqueous indirect photolysis of perfluoroalkanesulfonamides via reaction with •OH. All compounds shown were observed, with the exception of the aldehydes depicted in brackets. Formation of “other products” (unknown) indicates that an incomplete fluorine mass balance was measured. Multistep pathways are shown in more detail in Figures 3 and 4. 3664

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 10, 2009

TABLE 1. Observed Pseudo-First-Order Decay Rate Constant, k, and Products for Indirect Photolysis of Selected PFCs under Laboratory Conditions (10 mM Aqueous H2O2 and 765 Wm-2)

perfluorochemicala

k ( 95% confidence intervalb [h-1]

N-EtFOSE N-EtFOSAA N-MeFOSAA N-EtFOSA FOSAA FOSA PFOA

0.40 ( 0.06 0.12 ( 0.02 0.16 ( 0.02 0.19 ( 0.01 0.13 ( 0.02 no significant decay no significant decay

half-lifeb, t1/2 (95% confidence interval) [h] 1.7 (1.5, 2.0) 5.6 (4.7, 6.9) 4.4 (3.9, 4.9) 3.6 (3.4, 3.8) 5.2 (4.6, 6.0)

observed products N-EtFOSAA, N-EtFOSA, FOSAA, FOSA, PFOA N-EtFOSA, FOSAA, FOSA, PFOA FOSAA, FOSA, PFOA FOSAA, FOSA, PFOA FOSA, PFOA none none

fluorine mass balancec (range) 68%(24%) 28%(12%) 62% 93%(39%) 97%

a Irradiation experiments: N-EtFOSAA, n ) 4; N-EtFOSE, N-EtFOSA, FOSA, n ) 2; N-MeFOSAA, FOSAA, PFOA, n ) 1. Rate constants and half-lives are presented for comparative purposes, as the experimental concentration of H2O2 used produces a high (not environmentally relevant) concentration of hydroxyl radicals. c Mass balance at final time point.

b

FIGURE 3. (a) Proposed reaction mechanism for N-dealkylation of a perfluoroalkanesulfonamide via reaction with •OH. (b) Proposed reaction mechanism for oxidation of N-EtFOSAA to FOSAA via an iminium ion. (c) Proposed reaction mechanism for oxidation of N-EtFOSAA to N-EtFOSA via an iminium ion. source of •OH. The second-order rate constant was calculated using eq 7 from Boreen et al. (32), which applies to pseudo first order reaction systems evaluated under identical conditions. The approach used in the present study required a higher concentration of reference (∼0.15 mM acetophenone) than substrate (∼0.03 µM N-EtFOSAA), because acetophenone had a much higher detection limit. In addition, acetophenone reacted more quickly (t1/2 ) 0.7 h) than N-EtFOSAA (t1/2 ) 6 h). For these reasons, the [•OH ]SS in the reference reaction mixture may have been lower than in the substrate mixture. Therefore, the second-order rate constant of N-EtFOSAA calculated using the acetophenone decay rate represents an upper limit. Because the acetophenone solution colored over time, initial rates were used in order to avoid calculation error arising from light attenuation (and thus decreased •OH

production). Dark controls for both N-EtFOSAA and acetophenone with H2O2 were also irradiated and showed no decay.

Results and Discussion UV-Visible Absorbance. The absorbance of the PFCs tested in the present study was measured to confirm the lack of light absorption in the range of natural sunlight (Figure 1). Since no light is absorbed in this region, direct photolysis in the natural environment is not expected. The perfluoroalkanesulfonamides and PFOA absorb UV light, with maximum absorption near ∼204 nm. PFOA and other PFCs have been observed to degrade to F-, CO2, and short-chain PFCs when irradiated with UV light in wastewater treatment studies (16, 23). Indirect Photolysis. Irradiation of the perfluoroalkanesulfonamides N-EtFOSE, N-EtFOSAA, N-MeFOSAA, N-EtVOL. 43, NO. 10, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

3665

FIGURE 4. (a) Proposed reaction mechanism for oxidation of N-EtFOSE to N-EtFOSAA via reaction with •OH. (b) Proposed reaction mechanism for transformation of FOSAA to PFOA via reaction with •OH. FOSA, and FOSAA (see Figure 2 for structures) in the presence of H2O2 resulted in significant degradation relative to controls (see SI Figures S1-S4), indicating that reaction with •OH photochemically produced in the aquatic environment will result in the indirect photolysis of these compounds. The observed decay rate constants, products, and fluorine mass balances are reported in Table 1and were determined from independent irradiations of each starting compound. In addition to the formation of PFOA and FOSA observed in all cases, irradiation of N-EtFOSE in the presence of H2O2 resulted in N-EtFOSAA, N-EtFOSA, and FOSAA formation; N-EtFOSAA irradiation produced N-EtFOSA and FOSAA, and N-EtFOSA irradiation produced FOSAA. Irradiation of FOSAA produced solely PFOA and FOSA. Analogous to N-EtFOSAA, the irradiation of N-MeFOSAA produced FOSAA, PFOA, and FOSA (N-MeFOSA was likely produced but was not monitored due to the lack of an analytical standard). PFOA and FOSA appeared to be the final degradation products, as neither further degraded nor were products observed during independent irradiations. The resistance of PFOA to •OH initiated oxidation is consistent with the findings of Hatfield (26). Gauthier and Mabury (25) also observed PFOA as the major product of the indirect photolysis of 8:2 FtOH.

The formation of PFOS was not observed, which differs from the findings of D’eon et al. (12) for the gas-phase atmospheric reaction of N-methyl perfluorobutane sulfonamidoethanol (N-MeFBSE) with •OH, possibly due to the use of solvent (water) in the present study. N-MeFBSE is analogous to the starting compound used in the present study, N-EtFOSE, with a shorter perfluorocarbon chain and methyl- rather than ethyl- substitution of the sulfonamide. D’eon et al. (12) observed transformation to PFBA (analogous to PFOA; major pathway) as well as PFBS (analogous to PFOS; minor pathway). In agreement with our findings, Hatfield (28) did not observe PFOS formation following aqueous irradiation of N-EtFOSE and H2O2. Product Pathway. The products observed during each independent irradiation along with expected radical chemistry were used to generate the proposed pathway given in Figure 2. Because complete mass balances were not observed in all cases (likely due to an abundance of unspecific radical reactions), the formation of “other products” is depicted in the pathway. Mechanistic explanations for several of the reactions in Figure 2 are given in Figures 3 and 4. In several cases, loss of an R-group (ethanol, ethyl, or acetic acid) bonded to the

TABLE 2. Bimolecular Reaction Rate Constants for Reaction of •OH and Several Organic Contaminants, With Corresponding Half-Lives Calculated for a Range of Environmentally Relevant Steady State •OH Concentrations compound N -EtFOSAAa herbicidesb (triclopyr, dicamba, thiobencarb, chlorsulfuron, sulfometuron, 2,4-D, mecoprop, oxadiazon, bensulfuron methyl) sulfa drugsc (sulfamethazine, sulfamerazine, sulfadiazine, sulfachloropyridazine, and sulfadimethoxine) chlorinated phenolsd (2,4,6-trichlorophenol, 2,4-dichlorophenol, and 2-chlorophenol)

half-life (d) where [•OH] ) 10-18 M

half-life (d) where [•OH] ) 10-14 M

(1.7 ( 0.7) × 109 1.2 × 109 to 7.2 × 109

4.7 × 103 6.7 × 103 to 1.1 × 103

0.47 0.67 to 0.11

(3.7 ( 0.5) to (6.1 ( 0.6) × 109

2.2 × 103 to 1.3 × 103

0.22 to 0.13

1.3 × 103 to 9.8 × 102

0.13 to 0.10

kOH (M-1s-1)

6.3 × 109 to 8.2 × 109

a Upper limit of kOH, measured in present study. b kOH measured by Armbrust, 2000. 2005. d kOH measured by Farhataziz and Ross, 1977, and Tang and Huang, 1996.

3666

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 10, 2009

c

kOH measured by Boreen et al.,

nitrogen was observed upon reaction with •OH. As noted by D’eon et al. (12), N-dealkylation reactions are common in aqueous systems and may occur via the mechanism proposed by Tauber and von Sonntag (33) for atrazine, involving peroxy radical formation and loss of the hydroperoxyl radical HO2•. For example, N-EtFOSA may undergo N-deethylation via this mechanism (Figure 3a). When the nitrogen is not bound to an available hydrogen, we propose that N-dealkylation may instead occur by formation and hydrolysis of an iminium ion (34) as shown for formation of FOSAA (Figure 3b) and N-EtFOSA (Figure 3c) from N-EtFOSAA. Alternatively, oxidation of the R-group may occur by •OH attack via the mechanism shown for N-EtFOSE in Figure 4a (22, 35), in which N-EtFOSAA is produced. The oxidation of the aldehyde intermediate to the acid may occur spontaneously in oxygenated waters or be •OH-mediated. The perfluorooctanesulfonamides may also proceed directly to the endproduct PFOA via •OH attack at the sulfur atom, as proposed by D’eon et al. (12) and illustrated for FOSAA in Figure 4b. The mechanism involves the reaction of a perfluorinated peroxy radical with another peroxy radical (RHO2•), forming a tetroxide (22, 36) which may produce an aldehyde and a perfluorinated alcohol. The alcohol can lose HF to generate an acid fluoride, which produces a perfluorinated carboxylic acid upon contact with water (12). Hydroxyl Radical (•OH) Reaction. The measured upper limit for the bimolecular reaction rate constant for reaction of •OH and N-EtFOSAA (SI Figure S2) was (1.7 ( 0.7) × 109 M-1s-1, near the diffusion-controlled limit of ca. 1010 M-1s-1 and similar to rate constants measured for other organic compounds (Table 2). The rate constant is useful for estimating degradation rates expected for PFCs in sunlit aquatic environments. Extrapolation to Environmental Conditions. The rate of indirect photolysis of PFCs under environmental conditions will depend upon [•OH]SS (or other PPRIs), which ranges 10-18 to 10-14 M in freshwaters (17, 18). In their study of the aqueous, indirect photolysis of 8:2 FtOH, Gauthier and Mabury (25) measured a decay rate constant of 0.84 h-1 (t1/2 ) 0.83 ( 0.2 h) in 10 mM H2O2 compared to 0.007 h-1 (t1/2 ) 93 ( 10 h) in natural water from Lake Ontario. Perfluoroalkanesulfonamide decay rates measured in the present study in 10 mM H2O2 were slower than that of 8:2 FtOH by 50-90%. Thus for the natural water tested by Gauthier and Mabury (2005), longer half-lives (weeks to months) are expected for perfluoroalkanesulfonamides. A better estimate of environmental degradation rate may be calculated given the environmental concentration of •OH and the bimolecular reaction rate constant, as was measured for N-EtFOSAA in the present study. Table 2 reports the halflives calculated for N-EtFOSAA over a range of environmentally relevant values of [•OH ]SS (17, 18), resulting in removal rates that are negligible to half-lives of less than a day. For comparison, bimolecular reaction rate constants and expected half-lives are also shown for several other organic contaminants (37-40). The rate constants are similar, suggesting that the perfluorooctyl group does not appear to influence the reaction rate significantly. In sunlit environments, indirect photolysis of PFCs is likely to be important in the determination of their environmental fate given the slow rates expected for biotransformation and weak sorption. Photolytic conversion of PFCs into refractory perfluorinated acids, mainly PFOA, could mean that a significant fraction of these compounds will accumulate in the world’s oceans. Globally, transport to the deep ocean and sediment burial are considered the only significant removal mechanisms for PFOA (41).

Acknowledgments Portions of this research were funded by the Santa Clara Valley Water District (San Jose, California), the California Department of Water Resources, the WateReuse Foundation, and the National Science Foundation Graduate Research Fellowship program. Any opinions, findings, conclusions, and recommendations expressed in this publication are those of the authors and do not necessarily reflect the views of the sponsoring agencies.

Supporting Information Available Analytical conditions and figures showing reaction of N-EtFOSE, N-EtFOSA, N-EtFOSAA, N-MeFOSAA, FOSAA, FOSA, and PFOA over time during indirect photolysis experiments are presented, including product formation, mass balances, and controls. This material is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited (1) Schultz, M. M.; Barofsky, D. F.; Field, J. A. Fluorinated alkyl surfactants. Environ. Eng. Sci. 2003, 20 (5), 487–501. (2) Schultz, M. M.; Barofsky, D. F.; Field, J. A. Quantitative determination of fluorinated alkyl substances by large-volumeinjection liquid chromatography tandem mass spectrometrycharacterization of municipal wastewaters. Environ. Sci. Technol. 2006, 40 (1), 289–295. (3) Boulanger, B.; Vargo, J.; Schnoor, J. L.; Hornbuckle, K. C. Detection of perfluorooctane surfactants in Great Lakes water. Environ. Sci. Technol. 2004, 38 (15), 4064–4070. (4) Boulanger, B.; Vargo, J. D.; Schnoor, J. L.; Hornbuckle, K. C. Evaluation of perfluorooctane surfactants in a wastewater treatment system and in a commercial surface protection product. Environ. Sci. Technol. 2005, 39 (15), 5524–5530. (5) Ellis, D. A.; Martin, J. W.; De Silva, A. O.; Mabury, S. A.; Hurley, M. D.; Andersen, M. P. S.; Wallington, T. J. Degradation of fluorotelomer alcohols: A likely atmospheric source of perfluorinated carboxylic acids. Environ. Sci. Technol. 2004, 38, 3316– 3321. (6) Higgins, C. P.; Field, J. A.; Criddle, C. S.; Luthy, R. G. Quantitative determination of perfluorochemicals in sediments and domestic sludge. Environ. Sci. Technol. 2005, 39 (11), 3946–3956. (7) Giesy, J. P.; Kannan, K. Perfluorochemical surfactants in the environment. Environ. Sci. Technol. 2002, 36 (7), 146 A152 A. (8) Kannan, K.; Tao, L.; Sinclair, E.; Pastva, S. D.; Jude, D. J.; Giesy, J. P. Perfluorinated compounds in aquatic organisms at various trophic levels in a Great Lakes food chain. Arch. Environ. Contam. Toxicol. 2005, 48 (4), 559–566. (9) Hansen, K. J.; Clemen, L. A.; Ellefson, M. E.; Johnson, H. O. Compound-specific, quantitative characterization of organic fluorochemicals in biological matrices. Environ. Sci. Technol. 2001, 35 (4), 766–770. (10) Rhoads, K. R.; Janssen, E. M.-L.; Luthy, R. G.; Criddle, C. S. Aerobic biotransformation and fate of N-ethyl perfluorooctane sulfonamidoethanol (N-EtFOSE) in activated sludge. Environ. Sci. Technol. 2008, 42, 2873–2878. (11) Dinglasan, M. J. A.; Ye, Y.; Edwards, E. A.; Mabury, S. A. Fluorotelomer alcohol biodegradation yields poly- and perfluorinated acids. Environ. Sci. Technol. 2004, 38, 2857–2864. (12) D’eon, J.; Hurley, M. D.; Wallington, T. J.; Mabury, S. A. Atmospheric chemistry of N-methyl perfluorobutane sulfonamidoethanol, C4F9SO2N(CH3)CH2CH2OH: Kinetics and mechanism of reaction with OH. Environ. Sci. Technol. 2006, 40 (6), 1862–1868. (13) Martin, J. W.; Ellis, D. A.; Mabury, S. A.; Hurley, M. D.; Wallington, T. J. Atmospheric chemistry of perfluoroalkanesulfonamides: Kinetic and product studies of the OH radical and Cl atom initiated oxidation of N-ethyl perfluorobutanesulfonamide. Environ. Sci. Technol. 2006, 40 (3), 864–872. (14) Boutonnet, J. C.; Bingham, P.; Calamari, D.; Rooij, C. d.; Franklin, J.; Kawano, T.; Libre, J.-M.; McCul-loch, A.; Malinverno, G.; Odom, J. M.; Rusch, G. M.; Smythe, K.; Sobolev, I.; Thompson, R.; Tiedje, J. M. Environmental risk assessment of trifluoroacetic acid. Hum. Ecol. Risk Assess.: Int. J. 1999, 5 (1), 59–124. (15) Higgins, C. P.; Luthy, R. G. Sorption of perfluorinated surfactants on sediments. Environ. Sci. Technol. 2006, 40, 7251–7256. VOL. 43, NO. 10, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

3667

(16) 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. (17) Mabury, S. A.; Crosby, D. G. The relationship of hydroxyl reactivity to pesticide persistence. In Aquatic and Surface Photochemistry, Helz, G. R.; Zepp, R. G.; Crosby, D. G., Eds.; Lewis Publishers: Boca Raton, FL, 1994; pp149-161. (18) Lam, M. W.; Tantuco, K.; Mabury, S. A. Photofate: A new approach in accounting for the contribution of indirect photolysis of pesticides and pharmaceuticals in surface waters. Environ. Sci. Technol. 2003, 37 (5), 899–907. (19) Chin, M.; Wine, P. H. A temperature-dependent competitive kinetics study of the aqueous-phase reactions of OH radicals with formate, formic acid, acetate, acetic acid, and hydrated formaldehyde. In Aquatic and Surface Photochemistry, Helz, G. R.; Zepp, R. G.; Crosby, D. G., Eds.; Lewis Publishers: Boca Raton, FL, 1994; pp 85-96. (20) Bolton, J. R.; Cater, S. R. Homogenous photodegradation of pollutants in contaminated water: An introduction. In Aquatic and Surface Photochemistry, Helz, G. R.; Zepp, R. G.; Crosby, D. G., Eds.; Lewis Publishers: Boca Raton, Fl, 1994; pp 467-490. (21) Haag, W. R.; Yao, C. C. D. Rate constants for reaction of hydroxyl radicals with several drinking water contaminants. Environ. Sci. Technol. 1992, 26 (5), 1005–1013. (22) Larson, R. A.; Weber, E. J. Reaction mechanisms in environmental organic chemistry. CRC Press: Boca Raton, 1994. (23) Hori, H.; Yamamoto, A.; Koike, K.; Kutsuna, S.; Osaka, I.; Arakawa, R. Persulfate-induced photochemical decomposition of a fluorotelomer unsaturated carboxylic acid in water. Water Res. 2007, 41, 2962–2968. (24) Hori, H.; Yamamoto, A.; Hayakawa, E.; Taniyasu, S.; Yamashita, N.; Kutsuna, S.; Kitagawa, H.; Arakawa, R. Efficient decomposition of environmentally persistent perfluorocarboxylic acids by use of persulfate as a photochemical oxidant. Environ. Sci. Technol. 2005, 39, 2383–2388. (25) Gauthier, S.; Mabury, S. Aqueous photolysis of 8:2 fluorotelomer alcohol. Environ. Toxicol. Chem. 2005, 24 (8), 1837–1846. (26) Hatfield, T. L. Screening Studies on the Aqueous Photolytic Degradation of Perfluorooctanoic Acid (PFOA); 3M Environmental Laboratory: St. Paul, MN, 2001; part a. (27) Hatfield, T. L. Screening Studies on the Aqueous Photolytic Degradation of Perfluorooctanoic Acid (PFOS); 3M Environmental Laboratory: St. Paul, MN, 2001; part b.

3668

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 10, 2009

(28) Hatfield, T. L. Screening studies on the Aqueous Photolytic Degradation of 2-(N-Ethylperfluorooctanesulfonamido)-Ethyl Alcohol (N-EtFOSE Alcohol); 3M Environmental Laboratory: St. Paul, MN, 2001; part c. (29) Plumlee, M. H.; Larabee, J.; Reinhard, M. Perfluorochemicals in water reuse. Chemosphere 2008, 72, 1541–1547. (30) Renner, R. The long and the short of perfluorinated replacements. Environ. Sci. Technol. 2006, 40, 12A–13A. (31) Mukerjee, P.; Mysels, K. J. Critical Micelle Concentrations of Aqueous Surfactant Systems; National Standard Reference Data System, 1970. (32) Boreen, A. L.; Arnold, W. A.; McNeill, K. Photochemical fate of sulfa drugs in the aquatic environment: Sulfa drugs containing five-membered heterocyclic groups. Environ. Sci. Technol. 2004, 38 (14), 3933–3940. (33) Tauber, A.; von Sonntag, C. Products and kinetics of the OHradical-induced dealkylation of atrazine. Acta Hydrochim. Hydrobiol. 2000, 28, 15–23. (34) March, J., Advanced organic chemistry. 3 ed.; John Wiley & Sons, Inc.: 1985. (35) Bahnemann, D. Photocatalytic detoxification of polluted waters. In the Handbook of Environmental Chemistry: Environmental Photochemistry; Boule, P., Ed.; Springer: New York, 1999; Vol. 2L, pp 285-349. (36) Russell, G. A. Deuterium-isotope effects in the autoxidation of aralkyl hydrocarbons. Mechanism of the interaction of peroxy radicals. J. Am. Chem. Soc. 1957, 79 (14), 3871–3877. (37) Boreen, A. L.; Arnold, W. A.; McNeill, K. Triplet-sensitized photodegradation of sulfa drugs containing six-membered heterocyclic groups: Identification of an so2 extrusion photoproduct. Environ. Sci. Technol. 2005, 39 (10), 3630–3638. (38) Armbrust, K. L. Pesticide hydroxyl radical rate constants: Measurements and estimates of their importance in aquatic environments. Environ. Toxicol. Chem. 2000, 19 (9), 2175–2180. (39) Farhataziz, P. C.; Ross, A. B. Selected Specific Rates of Reactions of Transients from Water in Aqueous Solution, NSRDS-NBS59; National Bureau of Standards: Washington DC, 1977. (40) Tang, W. Z.; Huang, C. P. Effect of chlorine content of chlorinated phenols on their oxidation kinetics by Fenton’s reagent. Chemosphere 1996, 33 (8), 1621–1635. (41) Prevedouros, K.; Cousins, I. T.; Buck, R. C.; Korzeniowski, S. H. Sources, fate and transport of perfluorocarboxylates. Environ. Sci. Technol. 2006, 40 (1), 32–44.

ES803411W