Perfluorooctanesulfonate Degrades in a Laccase-Mediator System

Article Cite This: Environ. Sci. Technol. XXXX, XXX, XXX−XXX

pubs.acs.org/est

Perfluorooctanesulfonate Degrades in a Laccase-Mediator System Qi Luo,† Xiufen Yan,†,‡ Junhe Lu,§ and Qingguo Huang*,† †

Interdisciplinary Toxicology Program, Department of Crop and Soil Sciences, University of Georgia, Griffin, Georgia 30223, United States ‡ School of Environmental and Chemical Engineering, Jiangsu University of Science and Technology, Zhenjiang, Jiangsu 212003, China § Department of Environmental Science and Engineering, Nanjing Agricultural University, Nanjing, Jiangsu 210095, China Downloaded via KAOHSIUNG MEDICAL UNIV on August 27, 2018 at 12:08:17 (UTC). See https://pubs.acs.org/sharingguidelines for options on how to legitimately share published articles.

S Supporting Information *

ABSTRACT: Perfluorooctanesulfonate (PFOS) is a compound that has wide applications with extreme persistence in the environment and the potential to bioaccumulate, and could induce adverse effects to ecosystems. We investigated the degradation of PFOS by laccase-induced enzyme catalyzed oxidative humification reactions (ECOHRs) using 1-hydroxybenzotriazole (HBT) as a mediator. Approximately 59% of PFOS was transformed over 162 days of incubation, and the reaction appeared to follow a pseudo-first-order model with reaction rate constant of 0.0066/d (r2 = 0.87) under one condition tested. Using differential absorption spectra and theoretical simulation, we elucidated the interaction between Cu2+/Mg2+ and PFOS, and proposed that Cu2+ and Mg2+ could serve as a bridge to bring the negatively charged PFOS and laccase to proximity, thus increasing the chance of radicals that are released from laccase to reach and react with PFOS. In addition, density functional theory modeling showed that PFOS complexation to the metal ions could unlock its helical configuration and decrease the C−C bond energy of PFOS. These changes allow the attack of PFOS C−C backbone by radicals to become easier. On the basis of products identification, we proposed that direct attack of PFOS by the HBT radical initiated the free radical chain reaction processes and led to the formation of fluoride and partially fluorinated compounds. These results suggest that ECOHR is a potential pathway by which PFOS could be degraded in the environment, and it may make a viable approach to remediate PFOS contamination via amendment of appropriate enzymes and mediators.

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INTRODUCTION Perfluoroalkyl acids (PFAAs) are a group of man-made organic chemicals that are extremely stable and have been widely used in nearly every aspect of our lives.1 A great deal of attention has been heightened toward these chemicals because of their global distribution, recalcitrance to natural degradation, and potential toxicity.2−5 The extreme thermal and chemical stability of PFAAs arises from the high energy carbon−fluorine bond (531.5 kJ/mol)6 and the strong shielding effect of the helical conformation of their molecular structures.7 PFAAs have been extensively applied in industry and consumer products, such as food packaging, firefighting foam formulation, and semiconductor production.8 Perfluorooctanoic acid (PFOA) and perfluorooctanesulfonate (PFOS) are the two PFAAs that are most frequently detected in the environment and their co-occurrence has been found in groundwater plumes of substantial sizes.9,10 The major concerns over PFOA and PFOS are their ability to bioaccumulate and further induce a variety of undesirable effects in animals including carcinogenesis, infertility, endocrine disruption, and immunotoxicity.11,12 The U.S. Environ© XXXX American Chemical Society

mental Protection Agency (EPA) has classified PFOA and PFOS as emerging contaminants of concern.13 The unique structure of PFOS impedes it from being degraded by traditional water treatment processes and utilized by microbes. There has not been a report to date indicating biotic or abiotic degradation of PFOS under mild reaction conditions,14 while numerous studies have aimed to develop means to degrade PFOS for treatment and remediation purposes.15 Sonochemical oxidation,16 UV-activated photoreduction,17 photolysis,18 and electrolysis19 are proven to be able to decompose PFOS on a laboratory scale. Reductants such as zero-valent iron20 and vitamin B1221 have been used to reduce PFOS. However, most of these reaction mechanisms are not expected to degrade PFOS under natural conditions, and their applications for treatment or remediation are often Received: February 12, 2018 Revised: August 2, 2018 Accepted: August 3, 2018

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DOI: 10.1021/acs.est.8b00839 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

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Environmental Science & Technology

dimethoxyphenol (DMP), and 1-hydroxybenzotriazole (HBT) were from Sigma-Aldrich (St. Louis, MO). 5Diisopropoxy-phosphoryl-5-methyl-1-pyrroline-N-oxide (DIPPMPO) was purchased from Enzo Life Sciences (Farmingdale, NY). Perfluoroctanesulfonate (PFOS) was purchased from Inpofine Chemical Company (Hillsborough, NJ). Other PFAAs and surrogate standard sodium perfluoro-1-[13C8]octanesulfonate (M8PFOS) were obtained from Wellington Laboratories (Ontario, Canada) (see Table S5 for a full list). The cupric/magnesium sulfates were obtained from Fisher Scientific (Pittsburgh, PA). All HPLC-grade organic solvents including acetonitrile, methanol, and dichloromethane were also purchased from Fisher Scientific (Pittsburgh, PA). Milli-Q water (18.2 MΩ/cm resistivity) was prepared using the Nanopure Barnstead purification system (Thermo Scientific, San Jose, USA). Experimental Setup. The experiment was carried out in a solution containing 1.0 μM (0.5 mg/L) PFOS and 10 mM Cu2+ (160 mg/L of CuSO4) or Mg2+ (120 mg/L of MgSO4). Appropriate amounts of laccase and HBT were added to different reactors every 6 days to induce ECOHRs. We chose to add Cu2+ or Mg2+, because it was found in our earlier study that certain cations played an essential role in PFOA degradation by ECOHRs.32 The working solution was prepared by dissolving 1.0 μM PFOS in 10 mM CuSO4 or MgSO4 solutions. The resulting pH values of the working solution were 4.9 and 6.5 for the Cu2+ and Mg2+ solutions, respectively. Laccase activity assay was reported by Park et al.38 One unit of laccase activity is defined as the amount of enzyme that causes one unit change in absorbance at 468 nm per minute of a DMP solution at pH 3.8 in a 1 cm light path cuvette (a description is provided in the Supporting Information).39 Seventy-two reactors were prepared for each cation group. Each reactor included 10 mL of working solution, 60 μL of 167 U/mL aqueous laccase stock solution (1 U/mL), and 20 μL of 10 mM HBT dissolved in acetonitrile (20 μM HBT: named 1−20 treatment) or 20 μL of acetonitrile (0 μM HBT: named 1−0 treatment). The reactors with the addition of 60 μL of Mill-Q water and 20 μL of acetonitrile with no laccase and HBT were designated as 0−0 controls. Degradation experiments were conducted using time-sequenced, multiple-addition scheme at 22 °C with continuously shaking at 120 rpm in an incubator (Innova 42, New Brunswick Scientific). The whole experiment lasted for 162 days. Every 6 days, the reactors were dosed with the same amount of freshly prepared laccase solution (1 U/mL) and HBT in acetonitrile (20 μM HBT) or acetonitrile. At selected times, a set of nine reactors including triplicates from 1−20, 1−0 treatments, and the 0−0 control were sacrificed and diluted with an appropriate volume of Milli-Q water to the final volume of 12.16 mL for all reactors. A 0.5 mL aliquot of the solution was withdrawn, adjusted to approximately pH 10.5 with 60 μL of 200 μM NaOH, and then spiked with 0.5 mL of 0.5 μM M8PFOS before solid phase extraction40 cleanup as reported in previous studies.32,41,42 The detail process was described in the Supporting Information. To determine whether BTNO is responsible for PFOS degradation, an additional experiment was conducted to compare PFOS degradation in the laccase−HBT system with or without DIPPMPO, a spin trap that can effectively scavenge BTNO.43 To this end, three groups of reactors were set up and processed using the same procedure as described above with 1 U/mL laccase and 20 μM HBT added every 6 days, except that

restricted by the requirements of special equipment and highenergy input.22 Enzyme-catalyzed oxidative humification reactions (ECOHRs) are critical in natural humification processes,23 catalyzed by extracellular phenoloxidases and peroxidases which are ubiquitously present in the environment,24 mediating degradation of lignocellulosic materials and formation of humic substances. It has been found that many persistent organic pollutants (POPs), such as polycyclic aromatic hydrocarbons (PAHs)25 and polychlorinated biphenyls (PCBs)26 can be degraded by ECOHRs and covalently bound to natural organic matter (NOM) to result in their detoxification.27 Laccases are a major group of phenoloxidases that able to catalyze ECOHRs and have wide industrial applications.28 A laccase usually contains four copper ions, namely, one T1, one T2, and two T3 copper centers. The T2 and T3 copper centers form a trinuclear copper cluster site which is involved in the binding of oxygen during its reduction to water. The T1 copper center is involved in the oxidation of the reducing substrate.29 Laccases are able to catalyze the single-electron oxidation of phenolic or anilinic compounds (i.e., mediators) to form active intermediates such as free radicals and quinones, which can further react with other recalcitrant organic matters that cannot be directly oxidized by laccase.30 Chemicals like 1-hydroxybenzotriazole (HBT), ferulic acid, and vanillin have been commonly used as mediators in the laccase-mediator system, since they have high reaction efficiency and low environmental impact and represent functionalities commonly present in NOM.31 The laccase-mediator system has been applied in delignifying pulp, decoloring denim, and detoxifying textile dyes.31 Since ECOHRs are effective in degrading POPs, they are potentially feasible for remediation of PFAAs as well. Recently, we have demonstrated that laccase is able to mediate PFOA degradation using HBT as the mediator.32 Certain divalent cations, such as Cu2+ and Mg2+, were found to play an important role in ECOHRs. For instance, it was demonstrated that divalent cations can bind to NOM and thus neutralize the negative charge of the molecules, making them more accessible to the laccase molecule that is also negatively charged.33 Divalent cations were also found to facilitate PFOA degradation during ECOHR by bridging the negatively charged PFOA and laccase to form complexes, thus allowing the attack of PFOA by the short-lived benzotriazole nitroxyl radicals (BTNO) that are generated from the reaction between laccase and HBT.34 It is of particular interest to verify and examine PFOS degradation during ECOHRs, because PFOS is more stable than PFOA,35 and technologies capable of degrading PFOA (i.e., persulfate oxidation36 and electrolysis on Ti/SnO2 −Sb electrode)37 were proven ineffective toward PFOS. We herein show data demonstrating PFOS degradation in the laccase−HBT system, the very first report of PFOS degradation using laccase. The reaction kinetics were evaluated, the factors controlling PFOS degradation rate examined, and the underlying mechanisms proposed based on the degradation products identified by high-resolution mass spectrometry (HRMS) and molecular modeling.

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MATERIALS AND METHODS Chemicals and Reagents. All chemicals used in the experiments were reagent-grade or higher and used as received. Laccase from Pleurotus ostreatus (EC 420−150−4), 2,6B


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Environmental Science & Technology 200 μM DIPPMPO was also added to one set of the reactors each time by adding 40 μL of a 500 mM DIPPMPO solution in Milli-Q water, and the same amount of water was added to the other reactors without DIPPMPO. Identification of Reaction Products. The concentration of fluoride was determined by ion chromatography (IC) at selected times, while other potential organic products were identified by HRMS at the end of incubation. To analyze fluoride concentration, a 3 mL aliquot was taken from each reactor and filtered through 0.2 μm acetate cellulose membrane prior to quantification by IC detailed in Supporting Information. The influence of the solution matrix on fluoride analysis by IC was evaluated via a standard addition test by spiking 15.8 μM fluoride to a reaction solution reflecting the total reagent additions during the 162 day ECOHR experiment. The recovery of fluoride was determined to be (95 ± 15)% (n = 3). More details can be found in the Supporting Information. For identification of other organic reaction products, a 10 mL aliquot sample was taken from each reactor and repetitively extracted with 1 mL of dichloromethane four times at the end of incubation. The extractants were then combined and reconstituted in 40 μL of methanol and analyzed using Orbitrap Elite ESI-HRMS (Thermo Scientific, San Jose, USA) in negative ionization mode by direct injection with full scan acquisition (details provided in the Supporting Information). In order to differentiate products that may be formed during electrospray ionization (ESI) process in HRMS measurements, additional ESI control samples were prepared by mixing laccase, HBT, and PFOS (see Supporting Information for detail) and then immediately extracted and subjected to the same HRMS analysis. The full scan spectra of the ECOHRs treatment samples were compared to the negative control (no PFOS but with laccase and HBT), 0−0 control and the additional ESI control samples. A peak with m/z value only detected in the ECOHRs treatment sample but not in any of the controls was considered as being possibly associated with a product. The molecular formula of possible products comprising C, H, O, N, S, and F was determined using Formula Finder in the Thermo Xcalibur program within a 5 ppm mass error tolerance, and the common rules including carbon−hydrogen ratio, nitrogen rule, and isotopic ratios were employed to exclude unreasonable formula. Targeted MS/MS analysis was then performed on the possible products using the Thermo Orbitrap Elite HRMS, and the molecular structure for each potential reaction product was deduced according to its ion fragmentation pattern. Differential UV−Vis Spectrometry. The interaction between cations and PFOS was investigated using the differential UV−vis spectrometry method.44 The same volume (1 mL) of CuSO4 or MgSO4 stock solutions with different concentrations were added to 4 mL of 300 μM PFOS solutions to obtain sequentially increasing Cu2+ or Mg2 concentrations from 2 to 300 μM. The resulting mixtures having cation/PFOS ratio increased from 1:150 to 1:1. Then, 15 mL of citric buffer was added to each mixture to maintain the pH at 4.9 for Cu2+ and 6.5 for Mg2+ system that was equal to the unbuffered solutions containing 10 mM cations. Additional reference solutions without cation or PFOS were prepared in the same manner. After sample preparation, all mixtures and reference solutions were allowed to equilibrate for 24 h. Absorbance spectra were recorded using Beckman DU800 spectrophotometer at a wavelength from 200 to 700 nm. A differential

absorbance spectrum (DAS) was calculated using the equation: ΔAλ ,DAS = Aλ ,mixture − Aλ ,cation

where Aλ,mixture and Aλ,cation are the absorbance at λ wavelength of the mixture solution, the corresponding reference cation solution without PFOS, respectively.

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RESULTS Rate of PFOS Degradation during ECOHRs. As shown in Figure 1, PFOS in both cation solutions was fairly stable in

Figure 1. Change of PFOS concentration in ECOHRs over time with the presence (A) 10 mM Cu2+ in the solution and (B) 10 mM Mg2+ in the solution. 0−0 control: control sample to which no laccase or HBT was added; 1−0 treatment: 1 U/mL laccase added every 6 days but no HBT; 1−20 treatment: 1 U/mL laccase and 20 μM HBT added every 6 days. Error bars represent standard deviations (n = 3).

the control reactor in which neither laccase nor HBT was present. The treatment with the repetitive addition of 1 U/mL laccase alone (named 1−0) followed a trend similar to that of the control, while continuous degradation of PFOS was observed in the treatment with the repetitive addition of 1 U/ mL laccase and 20 μM HBT (named 1−20 treatment) for both cation solutions. However, the extent of PFOS degradation was found to be greater with Cu2+ than Mg2+: 59.0 ± 7.42% of PFOS was degraded in Cu2+ solution and 34.5 ± 7.27% in Mg2+ solution after 162 days of incubation. The decomposition of PFOS in both 1−20 treatments followed the pseudo-first-order kinetics. The reaction rate constants (k) C


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Environmental Science & Technology were 0.0066/d (r2 = 0.87) and 0.0021/d (r2 = 0.86) for the Cu2+ and Mg2+ solutions, respectively (Figure S1). Role of Benzotriazole Nitroxyl (BTNO) Radical in PFOS Degradation. We hypothesize that BTNO radicals generated by the laccase−HBT reaction are able to attack PFOS and lead to its degradation. In order to verify this, an experiment was carried out to repeat the 1−20 treatment in the 10 mM Cu2+ solution in the absence and presence of an efficient scavenger of BTNO radical, DIPPMPO.43 It was evident that PFOS degradation was considerably suppressed in the 1−20 treatment with DIPPMPO present in the solution, while significant reduction (43.7 ± 7.34%) of PFOS concentration was found in the corresponding 1−20 treatment without DIPPMPO over 108 days of incubation (Figure 2). This phenomenon strongly supports the hypothesis that BTNO radicals induced PFOS degradation.

Figure 2. Change of PFOS concentration in 10 mM Cu2+ solution during ECOHRs over time with the addition of DIPPMPO as a HBT radical scavenger. 0−0 control: control sample to which no laccase or HBT was added; 1−20 treatment: 1 U/mL laccase and 20 μM HBT added every 6 days; 1−20 RS treatment: 1 U/mL laccase, 20 μM HBT, and 200 μM DIPPMPO added every 6 days. Error bars represent standard deviations (n = 3).

Figure 3. UV−vis differential absorbance spectra (DAS) calculated according to the data recorded at varying cation concentrations (2.0, 20, 50, 100, 200, and 300 μM) while maintaining PFOS concentration at 300 μM. (a) Cu−PFOS DAS determined at pH 4.9; (b) Mg− PFOS DAS determined at pH 6.5.

systems, and the peak intensity decreases with the increase of Cu2+ and Mg2+ concentration. The peak may reflect the shift of electronic density in PFOS molecule by complexation with Cu2+ and Mg2+.44,46 Such a result confirms that Mg2+ and Cu2+can interact with PFOS as Fe3+ does.18 A peak at 277 nm with gradually decreasing intensity shows up only in the Cu− PFOS system but not in the Mg−PFOS system, while the intensity of the peak around 560 nm decreases in the Mg− PFOS system but not in the Cu−PFOS system, indicating that PFOS may interact differently with Mg2+ and Cu2+, therefore leading to the difference in PFOS degradation efficiency. An attempt was made to explore the complexation between the cations and linear PFOS using density functional theory (DFT) method. The molecular structures of PFOS anion and the PFOS−cation complexes were optimized by B3LYP method at 6-311+G(d,p) level (details are available in the Supporting Information). A comparison of the molecular geometries is provided in Figure 4 and the calculated structural parameters are listed in Tables S1−S3. The optimized structure of PFOS anion has a helical conformation (Figure 4a). After complexation to Cu2+ or Mg2+, the bond lengths in PFOS anion, including C−S and C−C, vary only slightly (Table S1), but its conformation changes considerably.

It is known that laccase catalyzes the oxidation of HBT to generate BTNO radical. The consumption of HBT during both 1−20 treatments shown in Figure 1 was documented (Figure S2). It was found that the total quantities of HBT converted in Mg2+ and Cu2+ systems were not significantly different (4.62 ± 0.90 μmole for Cu2+ and 5.21 ± 0.72 μmole for Mg2+) out of the total 5.40 μmole of HBT added to each system; however, the PFOS transformation efficiency was significantly higher in the Cu2+ system than in the Mg2+ system. Impact of Cations on PFOS Degradation by ECOHRs. A previous study reported that the formation of a complex between Fe3+ and PFOS enhanced the photodegradation of PFOS.18 A differential absorbance spectra (DAS) approach,44 a tool that is commonly used to probe the complexation between organic constituents and metal ions,45 was adopted to investigate possible interactions between PFOS and Cu2+ or Mg2+ in the ECOHR systems. The differential UV−vis absorbance spectra ranging from 220 to 620 nm (Figure 3) exhibit subtle but consistent changes when PFOS was mixed with Cu2+ or Mg2+ at increasing concentrations. As illustrated in Figure 3, a distinct positive peak is shown in the DAS at the wavelength of 313 nm for both Cu−PFOS and Mg−PFOS D


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Figure 4. Molecular structures in water optimized by B3LYP/6-311+G(d,p) method on (a) linear PFOS anion, (b) linear PFOS−Cu, and (c) linear PFOS−Mg. Gray, blue, yellow, and red colors represent carbon, fluorine, sulfur, and oxygen atoms, respectively.

Table 1. Partial Dihedral Angles of PFOS−Cu and PFOS−Mg Complexes in Water Optimized by B3LYP/6-311+G** Method. PFOS

PFOS−Cu

PFOS−Mg

dihedral angles

angles (deg)

angles (deg)

Δ

angles (deg)

Δ

S(26)C(8)C(1)C(5) C(5)C(1)C(8)C(7) S(26)C(8)C(6)C(4) S(26)C(8)C(3)C(2)

−109.45 −85.15 −57.41 −84.96

−104.81 −83.11 −52.84 −80.51

4.64 2.04 4.57 4.45

−106.78 −81.08 −54.44 −81.85

2.67 4.07 2.97 3.11

concentration in the 1−20 treatment sample at the end of incubation was 5.41 ± 0.57 μmol/L, significantly higher (more than 10-fold) than that in the corresponding 0−0 control sample measured at the same time (0.41 ± 0.12 μmol/L). Figure 5 demonstrates the change of fluorine mass distribution between dissolved fluoride and the fluorine remaining in PFOS during 1−20 treatment in 10 mM Cu2+ solution. As seen in Figure 5, the sum of the fluorine mass in the fluoride form and those remaining in PFOS accounted for approximately 73% of the total fluorine in the initial reaction system. The F− concentration increased with time at a rate of 0.617 μmol F−/day in the Cu2+ solution, while in the Mg2+ solution, the fluoride concentrations were only determined at the end of

Compared to the dihedral angels S(26)C(8)C(1)C(5) of PFOS (−109.45°), those of PFOS−Cu and PFOS−Mg decrease to −104.81 and −106.78°, respectively, while the other dihedral angles decrease as well (Table 1). This shows that complexation to Cu2+ and Mg2+ tends to slightly unlock the helical structure of PFOS, thus probably making the access by BTNO easier. Moreover, the DFT calculation of bond energy indicates that such complexation also reduce the C−C bond energy (Table S4). Major Reaction Products of PFOS by ECOHRs. Formation of fluoride is a key indicator of PFOS degradation. The concentrations of fluoride in selected 1−20 treatment and 0−0 control samples were monitored using IC. The fluoride E


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(Table 2). The MS/MS spectra of all tentative products are presented in Figures S4−S7, and their peak intensities were summarized in Table S6. For those products containing the 32S element such as products 1, 2, 7, 12, 13, and 16, their corresponding 34S isotope intensities were also used to verify the product structures. Product 15 was detected in both Cu2+ and Mg2+ systems. None of the reaction products identified were shorter-chain perfluorocarboxylic acids (PFCAs) reported for sonochemical,48 photodecomposition,18 or electrochemistry49 PFOS degradation technologies in earlier studies. It was found that all of the tentative products were partially fluorinated compounds (Table 2), which are similar to those found in our earlier studies of PFOA degradation during ECOHRs. These products can be categorized into three different groups by their structural features: (i) ones with a shorter-carbon chain perfluoroalkyl end-like products 3, 6, 7, 10, 11, and 15; (ii) ones with a perfluoroalkyl sulfonic acid end such as products 12, 13, and 16; and (iii) ones containing multiple nonfluoro moieties embedded with perfluoro-moieties (products 1, 2, 4, 5, 8, 9, and 14). It is worth noting that the molecular structures of products 9, 12, and 14 contained HBT moieties. This is the direct evidence confirming the reaction between BTNO and shorter-chain fluoroalkyl radicals. It should also be noted that the product identification was only performed at the end of 162 days of incubation, representing a snapshot of a relatively late stage of the incubation.

Figure 5. Change of fluorine mass distribution during 1−20 treatment in 10 mM Cu2+ solution. The normalized fluorine mass is calculated by the mass of the fluoride or the fluorine in remaining PFOS divided by the fluorine mass contained in PFOS prior to treatment. Initial PFOS concentration 1.0 μM, with 1 U/mL laccase and 20 μM HBT added every 6 days as indicated by the arrows. The result of PFOS degradation is shown in Figure 1a. Error bars represent standard deviations (n = 3).

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incubation, as 3.18 ± 0.16 and 0.17 ± 0.12 μmol/L, respectively, in 1−20 treatment and 0−0 control samples. The fluoride release ratio (R) was calculated from the C − equation R = C ×F 17 × 100% where CF− is the concentration

DISCUSSION The apparent difference in PFOS degradation rate during ECOHR in the Cu2+ and Mg2+ solutions, 0.0066/d and 0.0021/d, respectively (Figure S1), did not seem to be attributable to the influence of metal ions on the HBT reaction in generation of BNTO radicals, as slightly more HBT was converted in Mg2+ solution than in Cu2+ solution and HBT conversion was nearly complete in both systems (Figure S2). The difference may however arise from the different impact on PFOS structure when it complexing with different cations, evidenced by the DAS displayed in Figure 3 and by DFT simulation shown in Figure 4. Consistent with PFOS degradation rate, the release of fluoride in the treatment samples tended to be slower in the Mg2+ system than Cu2+ system as well, with the fluoride concentration determined at 3.18 and 5.41 μmol/L, respectively. As such, the defluorination ratios were similar in both treatment systems, 47.1 and 47.4% for Mg2+ and Cu2+ solutions, respectively, which suggests PFOS undergoes similar degradation mechanisms albeit at different rates in the two systems. It is noticeable in Figure 1a that PFOS degradation in Cu2+ solution was faster during the first 48 days, reaching 27% removal in the 1−20 treatment, but slowed down from day 48 to 84, and then picked up again afterward, leading to 59% removal after 162 days. Such a trend mirrors that of fluoride release as seen in Figure 5. The cause of such a trend of PFOS degradation in three phases is unknown and warrants further investigation. There might be certain products accumulating during the first phase that competed for ECOHRs and thus inhibited PFOS degradation during the middle phase, and PFOS degradation rate resumed in the third phase after the products were further transformed during the second phase. The previously proposed PFOS decomposition pathway usually starts with the dissociation of the sulfonic functional

0

of released fluoride, C0 is the concentration of PFOS at time zero, and 17 is the number of fluorine atoms contained in a PFOS molecule. The fluoride release ratios thus calculated were 31.8 and 18.7% for the 1−20 treatments in the Cu2+ and Mg2+ solution, respectively, at the end of 162-day treatment. Alternatively, the defluorination ratio (D) was also calculated C − by the equation D = (C − CF ) × 17 × 100%, where Ct is the 0

t

concentration of PFOS at time t.47 The defluorination ratios obtained by this equation reflect the extents of fluoride release relative to the total fluorine contained in the PFOS that had been transformed. They were 47.4 and 47.1% for the 1−20 treatments in the Cu2+ and Mg2+ solutions, respectively, at the end of 162 days of treatment. To identify possible products, we have extracted and analyzed the 1−20 treatment samples at the end of incubation. The workflow that we have used to screen possible products from ECOHRs is given in Figure S3. By comparing the highresolution mass spectra of the 1−20 treatment samples with the corresponding 0−0 control (PFOS only), negative control (no PFOS but with laccase and HBT), and the additional ESI controls (PFOS, laccase, and HBT added all at once and immediately extracted, see the Supporting Information for details), the m/z peaks corresponding to the molecular ions only present in the 1−20 treatment samples, but not in any of the control samples, were identified as possible product candidates. The element compositions of these candidates were then determined according to their accurate molecular weights given by HRMS (mass accuracy ≤ 5 ppm). The product candidates containing fluorine were further verified by MS/MS using the same HRMS instrument, and their possible structures were deduced from their fragment ion patterns F


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Table 2. Molecular Formulas, Theoretical and Measured Deprotonated Molecule Weight, Mass Error (ppm), and Possible Structures of PFOS Degradation Products from ECOHRs.

group, followed by stepwise unzipping of CF2 units from the perfluoroalkyl carbon backbone.48,50 Such a pathway would generate shorter-carbon chain PFCAs as degradation products.19 However, the reaction products identified in this study indicate that other reaction mechanism may be the major route for PFOS degradation during ECOHRs. The DAS results and DFT modeling show that metal ion−PFOS complexation may unlock the helical configuration, and decreases the C−C bond energy of PFOS. These changes allow the attack of PFOS C− C backbone by BTNO becoming easier. In addition, divalent metal cations can complex with both PFOS and laccase, bridging these two negatively charged species. Thus, BTNO had a greater chance to attack PFOS once it was generated and released from the laccase catalytic center.33 On the basis of the products identification and DFT computation information, a free radical chain reaction and cross-coupling mechanism could be a potential reaction mechanism for PFOS degradation during ECOHRs. First, HBT activated by laccase to form free radicals (BTNO). The free radicals could directly attack the C−C bonds in PFOS, leading to the formation of shorter-chain length perfluoroalkyl free radicals, with or without the sulfonic functionality. In the

meantime, other nonfluorinated organic compounds present in the reactor can also be converted to radicals by BTNO during free radical propagation process (eqs 2 and 3 in Figure 6). The perfluoroalkyl free radicals and the nonfluorinated radicals could couple to produce the degradation products (eqs 4−6 in Figure 6). Such a free radical chain reaction and coupling process may happen repetitively, leading to products with multiple nonfluorinated moieties and fluorinated moieties embedded with each other (eqs 7 and 8 in Figure 6). This study demonstrates PFOS degradation in the laccase− HBT reaction system, showing the effectiveness of ECOHRs on the degradation of perfluoroalkylsulfonates. ECOHRs induce PFOS degradation via a radical chain reaction by directly attacking the C−C bond of PFOS and generating the perfluoroalkyl or acid radicals followed by formation of partially fluorinated products via radical rearrangement and cross-coupling. Products formed during ECOHRs having less fluorine and more hydrogen atoms are expected to be less toxic and more available for microbial degradation.51 The ECOHR mechanism may be effective in natural water and soil systems to transform and incorporate PFOS into the natural organic matter, thus detoxifying and immobilizing PFOS. Moreover, it G


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Figure 6. Reaction mechanism of PFOS degradation during ECOHRs.

Spectrometry Core Facility for help with the high-resolution mass spectrometry analysis.

is possible that ECOHRs may be engineered by amending catalysts and mediators to serve as a remediation technique to degrade PFOS and other perfluorinated chemicals in the environment.

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ASSOCIATED CONTENT

* Supporting Information S

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.8b00839. Experimental details, rate model fits, HBT conversion data, identification workflow, high-resolution tandem mass spectra, bond lengths and angles, DFT bond energies, UPLC-MS/MS data, molecular formulas, and additional data from mass spectroscopy (PDF)

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REFERENCES

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AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]. Telephone: (770) 229-3302. Fax: (770) 412-4734. ORCID

Xiufen Yan: 0000-0001-9409-6629 Junhe Lu: 0000-0001-6068-5820 Qingguo Huang: 0000-0001-8863-9404 Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS The study was supported in part by U.S. Department of Defense SERDP ER-2127 and the U.S. Air Force Civil Engineering Center BAA Project FA8903-12-C-0005. We thank Hao Zhang and Shangtao Liang for help with the laboratory work and Dr. Dennis R. Phillips and Dr. Chau-wen Chou from the University of Georgia Proteomic and Mass H


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Perfluorooctanesulfonate Degrades in a Laccase-Mediator System

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