PFOA and PFOS Are Generated from Zwitterionic ... - ACS Publications

Letter Cite This: Environ. Sci. Technol. Lett. XXXX, XXX, XXX−XXX

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PFOA and PFOS Are Generated from Zwitterionic and Cationic Precursor Compounds During Water Disinfection with Chlorine or Ozone Feng Xiao,*,† Ryan A. Hanson,† Svetlana A. Golovko,‡ Mikhail Y. Golovko,‡ and William A. Arnold§ †

Department of Civil Engineering, University of North Dakota, 243 Centennial Drive, Stop 8115, Grand Forks, North Dakota 58202, United States ‡ Department of Biomedical Sciences, University of North Dakota, 1301 Columbia Road North, Stop 9037, Grand Forks, North Dakota 58202, United States § Department of Civil, Environmental, and Geo-Engineering, University of Minnesota, 500 Pillsbury Drive Southeast, Minneapolis, Minnesota 55455, United States S Supporting Information *

ABSTRACT: Perfluorooctanoate (PFOA) and perfluorooctanesulfonate (PFOS) are anionic organic pollutants, which are widespread in the environment. They have become a global concern due to their persistence in the environment as well as their toxicity and bioaccumulative properties. In this study, we demonstrate that PFOA, PFOS, or both are produced from a group of four zwitterionic/cationic polyfluoroalkyl amide (FA) and sulfonamide (FS) compounds during conventional drinking-water disinfection with chlorine or ozone. FA compounds were readily degraded by chlorine and converted primarily to PFOA, likely by a Hofmann-type rearrangement. FS compounds were much less reactive toward chlorine; the generation of PFOS from the FSs was not significant. All four FA and FS compounds were degraded rapidly during ozonation, generating PFOA, PFOS, and a number of infrequently reported products for which chemical structures were either confirmed or tentatively proposed using high-resolution mass spectrometry. FSs generated both PFOS and PFOA during ozonation with the yield of PFOA even higher than that from the FAs. The results of this study may provide important insight into the degradation mechanisms of FAs and FSs and shed light on their contribution to the secondary formation of PFOA and PFOS in natural and engineered systems.

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INTRODUCTION

underlying the secondary formation of PFOA/PFOS from precursor compounds in these systems are poorly understood. With improvement in identification methods, more than 400 emerging ionic PFASs have recently been identified in aqueous film-forming foams (AFFFs), commercial surfactants, and soil and water samples from AFFF-contaminated sites.25−32 Approximately 32% and 20% of these emerging PFASs are zwitterions and cations, respectively.32 A majority of zwitterionic and cationic PFASs are derivatives of polyfluoroalkyl amides (FAs) or polyfluoroalkyl sulfonamides (FSs). These FAs and FSs are structurally similar to PFOA and PFOS, except that the perfluoroalkyl chain is attached to a nonfluorinated moiety through an amide (FAs) or sulfonamide (FSs) group. The fate, transport, and ecological impacts of FAs and FSs are not well understood. In a recent study, Mejia-Avendano et al.33 investigated aerobic soil degradation of one cationic FA (perfluorooctaneamido ammonium salt, PFOAAmS) and one

Perfluorooctanoate (PFOA) and perfluorooctanesulfonate (PFOS), members of a group of compounds known as perand polyfluoroalkyl substances (PFASs), have been detected in drinking water in many locations.1−3 PFOA and PFOS are both resistant to degradation4−7 and difficult to remove during conventional water and wastewater treatment.5,7−9 The U.S. EPA recently released an updated drinking water advisory on PFOA and PFOS specifying a maximum combined level of 70 ng/L,10 making removal of PFOA/PFOS from drinking water11−16 and remediation of PFAS-contaminated sites a priority issue.17,18 The contamination of drinking water by PFOA or PFOS may result from both direct3,19 and indirect20 sources. Indirect sources are precursor compounds that contribute to secondary formation of PFOA/PFOS in natural and engineered systems. In a few survey studies,12,21,22 an increase in concentrations of PFOA/PFOS has been observed in the disinfection process in drinking-water treatment plants. The generation of PFOA/ PFOS has also been noted in biological wastewater treatment8,23 and soil remediation processes.24 The mechanisms © XXXX American Chemical Society

Received: May 17, 2018 Revised: May 23, 2018 Accepted: May 24, 2018

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DOI: 10.1021/acs.estlett.8b00266 Environ. Sci. Technol. Lett. XXXX, XXX, XXX−XXX

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Environmental Science & Technology Letters

a dosage of 40 mg/L, 20 min flocculation, 30 min settling, and filtration through a Whatman paper filter to remove remaining fine particles. The pretreated water was further disinfected either by breakpoint chlorination with sodium hypochlorite (Acros Organics) at a free chlorine dose of 5.1 mg/L after the breakpoint or by ozonation at 3.4 mg O3/L. The doses of disinfectants are within typical ranges used for conventional drinking-water disinfection (e.g., 1−6 mg/L of free chlorine and 1−5 mg/L of ozone38,39). The chlorination breakpoint was identified in a separate experiment. In certain experiments, the chlorinated water received additional or booster chlorination. Booster chlorination is commonly used to maintain sufficient residual chlorine in large distribution systems. Ozonation was conducted by continuously bubbling ozone gas into the pretreated water through an air stone diffuser for up to 90 min. Ozone was produced by a generator (model: Z-7G, capacity: 7 g O3/h; A2Z Ozone Inc., Louisville, KY) with oxygen as the source. The input ozone dose was determined in distilled (DI) water buffered with 1 × 10−3 mol/L sodium bicarbonate (pH 7.4−7.6). For initial experiments, chlorination/ozonation was performed on a mixture of all four FAs and FSs to provide a baseline for subsequent degradation experiments with an individual FA/FS. To identify the degradation products, experiments were also performed in buffered DI water. Chlorinated/ozonated samples were quenched with sodium thiosulfate at a molar dose five times that of the dosed oxidant and microfiltered (0.45 μm) for determination of PFAS concentrations. More details of pretreatment and chlorination/ozonation experiments are available in the Supporting Information (SI). Quantification of PFASs and identification of degradation products were carried out on a Waters Acquity ultrahigh pressure liquid chromatography (UPLC) system coupled with a Waters QToF-MS (Synapt G2-S, Waters Corporation, Milford, MA, USA). Detailed operating parameters can be found in a previous study31 and in the SI. The yield of PFOA/PFOS was calculated by

cationic FS (perfluorooctanesulfonamido ammonium salt, PFOSAmS), and observed significant generation of PFOA from the FA. The results33 suggest FAs can be important precursor compounds of PFOA. FAs and FSs have recently been found in AFFF-impacted drinking water sources.26,29,34 One study has investigated the removal of FAs and FSs by adsorption from water.34 The authors predicted that FAs and FSs will break through granular activated carbon cartridges before PFOA and PFOS,34 suggesting that FAs and FSs will not be easily removed by physicochemical approaches. The removal and transformation behaviors of FAs/FSs during oxidation treatment (e.g., chemical disinfection) remain unknown. FAs and FSs may transform to toxic products including PFOA and PFOS by reaction with chemical disinfectants, becoming a secondary source of exposure to PFOA/PFOS for AFFF-impacted communities. To fill this knowledge gap, we investigated the reactions between two conventional drinking-water disinfectants (chlorine and ozone) and four model FA and FS compounds, PFOAAmS ( ), PFOSAmS ( ), and their corresponding zwitterionic betaine counterpartsperfluorooc) and perfluorooctanetaneamido betaine (PFOAB, ). PFOAB and sulfonamido betaine (PFOSB, PFOSAmS have been observed in AFFF and commercial surfactant samples.27,29,31 PFOSAmS and PFOSB have recently been found in soil samples,28,30 and PFOAB has been detected in bottled water samples.35 Other FAs and FSs could also be present in the natural environment,26,29,34 but these four chemicals are among the very few FAs and FSs for which genuine standards are available. We hypothesized that PFOA ( ) and PFOS ( ) are generated from FAs and FSs under typical conditions used for water disinfection. We also investigated whether there were other intermediate products using a high-resolution quadrupole time-of-flight mass spectrometer (QToF-MS). Finally, we propose the pathways leading to formation of PFOA and PFOS. The results from this work may lead to development of effective treatment/ remediation strategies while limiting the unwanted secondary production of PFOA and PFOS from precursor compounds.

Yield =

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MATERIALS AND METHODS PFOAB (pKa1 = −0.48, pKa2 = 2.27), PFOSB (pKa1 = 1.91, pKa2 = 3.30), PFOAAmS (pKa = −0.53), and PFOSAmS (pKa = 3.28) were purchased from Fluobon Surfactant Institute (China) (purity: PFOAB, 95%; PFOSB, 95%; PFOAAmS, 98%; PFOSAmS, 98%). Their pKa values were estimated using freeware (Marvin 15.10.26, ChemAxon, Cambridge, MA).32,36 PFOS (potassium salt) (>98%) and PFOA (96%) were purchased from Sigma−Aldrich. Each of the PFASs was dissolved in a solution composed of 50% methanol and 50% water to a concentration of 1 × 10−3 mol/L. The stock solution was stored at 4 °C before use and reprepared every month. The test solution was surface water collected from the Red River (Grand Forks, ND) (dissolved organic carbon = ∼4 mg/ L; specific ultraviolet absorbance = ∼2 L/mg/cm; pH ∼7.9; turbidity = 3.6−6.8 NTU). The water was spiked with FA/FS to a known concentration, equilibrated for 48 h at room temperature (22 °C), and pretreated by processes simulating typical drinking-water physicochemical treatment processes.9,37 Pretreatment included 1 min coagulation with alum (Al2(SO4)3·18H2O; purity: 98.0−102.0%; Fisher Chemical) at

Δ[PFOA or PFOS]formation × 100% Δ[FA or FS]degradation × (− 1)

(1)

where Δ[PFOA or PFOS]formation and Δ[FA or FS]degradation are molar concentration changes of corresponding PFAS.

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RESULTS AND DISCUSSION As shown in SI Figure S1, conventional coagulation/ flocculation at an alum dose of 40−60 mg/L does not significantly remove cationic FA and FS compounds but can achieve up to a 20% removal of PFOAB (a zwitterionic FA). The electrostatic attraction9 between the negatively charged carboxyl group of PFOAB and the positively charged flocs37 may contribute to the removal. The chlorination results of a mixture of four FA and FS compounds in both the pretreated surface water and buffered DI water are given in SI Figures S2 and S3. The FAs (PFOAB and PFOAAmS) were degraded, and this process was accompanied by a significant generation of PFOA (SI Figures S2 and S3). Levels of the two FSs (PFOSB and PFOSAmS), on the other hand, were stable during chlorination. The decline in concentrations of FSs and the generation of PFOS were both insignificant (SI Figures S2 and S3). The yields of PFOS were only 0.2 and 0.5 mol % from B

DOI: 10.1021/acs.estlett.8b00266 Environ. Sci. Technol. Lett. XXXX, XXX, XXX−XXX

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Environmental Science & Technology Letters

Figure 1. Generation of PFOA from PFOAB and PFOAAmS during chlorination in the pretreated surface water and buffered DI water (pH ∼7.4). The ionic species of PFOAB/PFOAAmS/PFOA shown in (a) and (b) are estimated at circumneutral pH. The pseudo-first-order degradation rate constants (k) and half-lives (t1/2) in the buffered DI water are provided in (c) and (d). In both waters, the initial free chlorine dose after the breakpoint was 5.1 mg/L, and the residual chlorine was 1.1−1.5 mg/L after ∼24 h in the pretreated surface water and 2.1−3.0 mg/L in the buffered DI water after ∼110 h. In the booster chlorination process, the chlorinated surface water received an additional free chlorine dose at 5.1 mg/L, and the residual chlorine was 1.7−1.9 mg/L after ∼25 h.

Figure 2. Formation mechanisms of PFOA and PFOS from FAs (fluoroalkyl amides) and PFOSB (one FS) during chlorination and ozonation proposed in this study. The intermediate products of the proposed Hofmann-type rearrangement may be unstable and short-lived as they were not detected by HRMS. The •OH-induced PFOA formation pathway from PFOSB was proposed based on other studies.47−49 The PFOS formation pathway from PFOSB was proposed based on degradation products tentatively identified (compounds #5−#8) or confirmed (compound #9) with a standard by HRMS (see SI for UPLC chromatograms and QToF MS and MSE spectra).

chlorine = 2.1−2.3 mg/L), slowed down in the next 5 h (residual free chlorine = 1.5−1.8 mg/L), and then leveled off when the residual free chlorine declined further (Figures 1a and 1b). The yields of PFOA were 24−28 and 27−38 mol % from PFOAB and PFOAAmS, respectively, after 24−70 h of chlorination (Figure 1). Additional or booster chlorination of the chlorinated surface water led to further and greater

PFOSAmS and PFOSB, respectively, after 3-d chlorination in buffered DI water (SI Figure S4). We performed single-compound degradation experiments to further understand the formation mechanisms of PFOA from each of the FAs (Figure 1). In the pretreated surface water, the generation of PFOA from either PFOAB or PFOAAmS was rapid during the initial 2 h of chlorination (residual free C

DOI: 10.1021/acs.estlett.8b00266 Environ. Sci. Technol. Lett. XXXX, XXX, XXX−XXX

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Environmental Science & Technology Letters

Figure 3. Generation of PFOA and PFOS (closed symbols) from four FAs and FSs (open symbols) during ozonation of pretreated surface water. The ionic species of PFOSB/PFOSAmS/PFOS shown in (a) and (c) are estimated at circumneutral pH. The input ozone dose was 3.4 mg/L. Solid lines: first trial. Dashed lines: second trial.

PFOSAmS is stable during conventional water chlorination. A recent study found that PFOSAmS is also much more resistant to aerobic soil degradation than PFOAAmS.33 Taken together with the findings in that study,33 our results indicate that FSs may be in general less susceptible than FAs to relatively mild oxidation treatments (e.g., chlorination and aerobic soil biodegradation). While relatively stable during chlorination, FSs were quickly degraded during conventional ozonation, generating both PFOS and PFOA (Figure 3 and SI Figure S6). The yield of PFOS after ∼50 min of ozonation was 22−43 mol % from PFOSB and PFOSAmS. FAs (PFOAB and PFOAAmS) were also degraded, generating PFOA with a yield (∼2−6 mol %) lower than that from FSs (6−19 mol %) after ∼50 min of ozonation. We fit the initial 15 min of ozonation data in the pretreated surface water (Figure 3) to a pseudo-first-order kinetic model,50−52 which revealed that these compounds have a half-life of a few minutes during ozonation (t1/2,FSs = 3.2−4.5 min (95% CI = 3.2−7.6 min); t1/2,FAs = 6.5−7.9 min (95% CI = 4.7−10.5 min)). In addition to PFOA and PFOS, we also observed other degradation products from FSs and FAs (compounds #1−#3 and #5−#9 in SI Table S1), especially from PFOSB during ozonation. Genuine standards of most of the detected degradation products are not available. The structures in Table S1 are proposed based on the UPLC−QToF-MS/MSE spectra (SI Figures S7−S14). Many of them have also been

degradation of remaining PFOAB/PFOAAmS molecules and a higher yield of PFOA (46−60 mol %) than from the primary chlorination (Figure 1). The yield of PFOA was considerably high in the buffered DI water, 43−69 mol % from PFOAB and 74−95 mol % from PFOAAmS after 3-d chlorination (Figure 1 and SI Figure S5). The absence of competitive substances such as natural organic matter may lead to a high yield in the DI water. The degradation of FAs in the buffered DI water could be described by pseudo-first-order kinetics with respect to chlorine concentration. PFOAAmS has a shorter half-life (∼5.7−8.7 h) than PFOAB (∼9.1−13.5 h) (Figures 1c, 1d, and SI Figure S3). The different fates of FAs and FSs (e.g., PFOAB versus PFOSB) during chlorination are likely due to varying reactivity of amide and sulfonamide groups toward chlorine. We propose that the secondary amide group in the FAs undergoes a Hofmann-type rearrangement40−44 during chlorination (Figure 2). The first step is N-chlorination during which chlorine reacts with the amide group in a substitution reaction to result in an N-chloroamide. This step is followed by the cleavage of the carbon and nitrogen bond with the attack of OH− (or another nucleophile, such as OCl−), which separates the perfluoroalkyl chain from the nonfluorinated moiety and leads to the formation of PFOA (Figure 2). In some reports, 45,46 sulfonamides can also undergo a Hofmann rearrangement with the presence of special reagents or catalysts. Our results show, however, that the sulfonamide group in PFOSB and D

DOI: 10.1021/acs.estlett.8b00266 Environ. Sci. Technol. Lett. XXXX, XXX, XXX−XXX

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Environmental Science & Technology Letters tentatively identified (compounds #1, #3−#5, and #7 in SI Table S1) or confirmed (compound #9 in SI Table S1) during aerobic soil degradation of PFOAAmS and PFOSAmS.33 Compounds #2, #6, and #8 were detected for the first time in this study. Compound #9 (SI Table S1) is confirmed as perfluorooctane sulfonamide ( ) with an industrial standard provided by 3M (SI Figure S14). Most of the degradation products appear to be minor, with the MS spectral peak intensity less than 20% of that of PFOA generated from PFOAB/PFOAAmS during chlorination/ozonation or of that of PFOS generated from PFOSB and PFOSAmS during ozonation. The reaction between ozone and PFOSB/PFOSAmS generated both PFOS and PFOA (Figure 3). It is known that ozone degrades organic compounds in water through both direct ozone oxidation and indirect oxidation by means of •OH.53 Two pathways have been previously suggested for the production of PFOA from •OH-induced degradation of anionic and neutral FSs.47−49 In one pathway, the nonfluorinated chain of the FSs undergoes oxidation and Ndealkylation, and the fluorinated chain then transforms to perfluoroalkyl radical by S-dealkylation.49 In another pathway, FSs directly transform to perfluoroalkyl radical by S-dealkylation. The perfluoroalkyl radical further undergoes C radical oxidation and results in PFOA (Figure 2).49 The formation of PFOA during the ozonation of cationic and zwitterionic FSs (Figure 3) agrees well with the previous work.47−49 Notably, PFOS is not among the •OH-induced degradation products of FSs in the previous studies.47,48 Therefore, the observed generation of PFOS (Figure 3) may be a result of direct ozone oxidation of FSs (see Figure 2 for the proposed formation in a pathway from PFOSB). This hypothesis was supported by ozonation experiments in the presence of a strong •OH scavenger (5 mmol/L tertiary butanol54,55) where the yield of PFOS from PFOSB and PFOSAmS was enhanced to 51−63 mol % accompanied by a decreased yield of PFOA to 2−7 mol % (SI Figure S15). The •OH scavenger shifted the reaction to the direct ozonation pathway,55−58 leading to the increased yield of PFOS. The possibility, however, that PFOA is generated by direct ozonation of FSs cannot be excluded because PFOA was still produced in the presence of the •OH scavenger. The present study demonstrates the generation of PFOA, PFOS, or both from emerging FA and FS compounds during chemical disinfection with chlorine or ozone and provides novel information on the formation pathways. This study is highly relevant to AFFF-impacted communities. Our results indicate that the removal of these precursor compounds prior to chemical disinfection/oxidation would be necessary to limit the secondary formation of PFOA and PFOS in drinking-water treatment plants or during booster chlorination in distribution systems. Furthermore, along with a soil degradation study,33 the findings suggest that FAs and FSs recently identified in contaminated soils and natural water26,28−30 may have already contributed to the secondary generation of PFOA and PFOS in relevant natural and engineered processes.

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Surface water treatment details, UPLC−QToF-MS method, removal of four PFASs during coagulation/ flocculation (Figure S1), degradation of four PFASs during chlorination and ozonation (Figures S2−S6 and S15), identified degradation products (Table S1), UPLC−ESI-ToF MS or MSE chromatograms and spectra of degradation products and the standard of PFOSA (Figures S7−S14). (PDF)

AUTHOR INFORMATION

Corresponding Author

*Phone: (701)777-5150. Fax: (701)777-3782. E-mails: Feng. [email protected], [email protected]. ORCID

Feng Xiao: 0000-0001-5686-6055 William A. Arnold: 0000-0003-0814-5469 Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS This work was supported by the University of North Dakota Vice President of Research & Economic Development Faculty Early Career Award to F.X. (20622-4000-02545). The UPLC and the hybrid QToF HRMS systems were purchased under the NIH funded COBRE Mass Spec Core Facility Grant 5P30GM103329-05. The authors have no conflict of interest to declare.

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REFERENCES

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ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.estlett.8b00266. E

DOI: 10.1021/acs.estlett.8b00266 Environ. Sci. Technol. Lett. XXXX, XXX, XXX−XXX

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DOI: 10.1021/acs.estlett.8b00266 Environ. Sci. Technol. Lett. XXXX, XXX, XXX−XXX

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DOI: 10.1021/acs.estlett.8b00266 Environ. Sci. Technol. Lett. XXXX, XXX, XXX−XXX