Emerging investigator series: rapid defluorination of 22 per- and polyfluoroalkyl substances in water using sulfite irradiated by medium-pressure UV†

Ibrahim Abusallout , Junli Wang and David Hanigan *
Department of Civil and Environmental Engineering, University of Nevada, Reno, Nevada 89557-0258, USA. E-mail: dhanigan@unr.edu; Tel: +775 682 7517

First published on 25th June 2021

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Water impact A high flux medium-pressure UV lamp coupled with dissolved sulfite defluorinated per- and polyfluoroalkyl substances faster than low-pressure lamps and was less affected by operational conditions. This will allow operation with more variable water chemistry, such as groundwater treatment, but comes at the cost of high energy use.

1. Introduction

Research has recently focused on the development of treatment methods for per- and polyfluoroalkyl substances (PFAS) in water due to their occurrence in drinking water, wastewater and groundwater.^1–4 PFAS are highly recalcitrant in the environment and poorly degraded by conventional water treatment processes,^5–8 which is attributable to the strength of the C–F bond. Consequently, PFAS are broadly distributed in multiple environmental compartments and have been detected in human serum at microgram per liter levels.^9–15 Toxicological studies have shown that some PFAS are associated with adverse human health outcomes.¹⁶ The most common PFAS present in water are perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS), but many different PFAS species (e.g., varying fluoroalkyl chain length, terminal functional group) have been detected in water due to contamination from highly concentrated PFAS sources in many industrial and commercial applications.^17,18 Therefore, the need for an effective water treatment technology which removes or destroys PFAS is crucial to protect the public from health risks associated with PFAS exposure.

Many treatment methods have been examined for PFAS removal from water including sorption,^19–24 filtration,^25–27 chemical oxidation,^28–31 electrochemical oxidation/reduction,^32–35 sonolysis^36–39 and biodegradation.^40–43 However, disadvantages surrounding these methods have hindered their applicability including poor selectivity, formation of toxic byproducts, and complex operation. Reduction via irradiation of sulfite and production of hydrated electrons (e_aq⁻) has been shown to defluorinate PFAS at bench-scale but at relatively slow degradation rates, limiting its scalability.⁶⁴ Further examination of reduction methods may lead to scalable PFAS treatment technology which is field-ready.

UV/sulfite generates e_aq⁻ through photoionization of sulfite (SO₃²⁻) (eqn (1)). e_aq⁻ have standard reduction potential of −2.9 V,⁷¹ and are capable of cleaving C–F bonds in PFAS, resulting in fluoride (F⁻) ions. Other electron donors have also been examined including inorganic compounds (e.g. iodide, ferrous minerals^46–49), organic compounds (e.g. amino acids⁵⁰) and indoles,⁵¹ but sulfite has the advantage of forming the relatively innocuous products sulfate and bisulfite.⁵² Several other factors in addition to electron donors control e_aq⁻ production and longevity including pH, dissolved oxygen (O₂) and other natural e_aq⁻ scavengers. pH impacts the formation of sulfite intermediates (e.g., HSO₃⁻ and S₂O₆²⁻)⁵³ and O₂ consumes e_aq⁻, forming oxidative radicals including superoxide.⁴⁴ Dissolved organic matter (DOC) and nitrate also scavenge e_aq⁻, thus reducing the dehalogenation of the targeted compound by UV/sulfite.^55,79

SO32− + hv → SO3˙− + eaq−	(1)

Low pressure (LP) UV lamps produce 254 nm photons and are the most commonly studied lamp because of the overlap between the produced photons and the absorbance spectra of sulfite.⁴⁵ However, in research conducted by Bentel et al., LP UV/sulfite treatment of eight different perfluorocarboxylic acids (PFCAs, C₂ to C₉) resulted in 49 to 94% defluorination after 24 h at pH 9.5 and 79 to 100% after 8 h at pH 12.⁵³ Others have shown that greater photon flux produced by polychromatic high-pressure (HP) UV lamps results in rapid defluorination of PFOA and PFOS, ∼45% and ∼30% after 10 min of UV/sulfite, respectively.^55,56 Although these results are promising for PFOA and PFOS destruction, HP lamps are not commonly used in water treatment due to high operational costs and less effective spectrum for specific application such as disinfection or advanced photochemical reactors.⁵⁷ Further, HP lamps consume more energy due to higher emission intensities than other UV lamps.^55,58 There is a need to examine other rapid PFAS defluorination methods with consideration of energy inputs and water treatment applications. No work has been conducted with medium-pressure (MP) lamps in presence of sulfite, which have emission spectra similar to HP lamps (200–600 nm) but with more energy focused in UVB and UVC range than HP lamps, which is needed for sulfite excitation and e_aq⁻ production.^57,59–63

To fill this knowledge gap, we investigated the defluorination rates of a diverse group of PFAS using MP UV/sulfite system. The effects of O₂ content, pH, sulfite concentration and aqueous matrix on PFAS defluorination rate were measured. This study provides a dataset of defluorination kinetics for 22 PFAS and a pathway towards full-scale implementation of UV/sulfite degradation systems.

2. Methods and materials

2.1 Chemicals and reagents

22 PFAS species were examined including 11 perfluorocarboxylic acids (PFCAs), 3 fluorotelomer alcohols (FTOHs), 2 perfluorosulfonic acids (PFSAs), 3 iodinated PFAS, 1 fluorotelomer acrylate, (perfluorohexyl)ethylene, and ammonium perfluoro-2-methyl-3-oxahexanoate (GenX). Details including supplier, CAS number and purity for each of the 22 PFAS are provided in Table S1 in the ESI.† Individual PFAS stocks were prepared in either ultrapure water or methanol (MeOH) (0.2%) at 0.5 g F per L and stored at 4 °C. Borate buffer (for near neutral samples), sodium hydroxide and nitric acid were used to adjust pH and purchased in addition to sodium sulfite (Na₂SO₃, 98%) from Fisher Scientific (Waltham, MA). All solutions used in this study were prepared with ultrapure water (≥18.2 MΩ cm) produced by a Thermo Fisher Genpure system.

2.2 Photochemical reactor

PFAS defluorination experiments were conducted using a high energy 450 W, MP mercury-vapor quartz lamp (121.92 mm L, 25 mm I.D., Product # 7825-340, Ace Glass Incorporated, Vineland, NJ). This lamp emits 40 to 48% of the total radiated energy in the UV range (200 to 400 nm), 40 to 43% in the visible range and 10 to 20% in the infrared (manufacturer provided emission spectrum, Fig. S2†) the lamp was inserted in a quartz cooling jacket with tap water as the cooling agent without recirculation. The temperature of the surface of the reactor was 20 °C throughout the experiments. Photolysis experiments were performed using 15 mL quartz tubes with an outside diameter of 1.8 cm and a length of 7 cm. The thickness of the tube wall was 0.1 cm. Tubes were capped to prevent volatilization or evaporation during experiments. Samples were placed 3 cm from the center of the MP lamp. This distance was recommended by the manufacturer of the UV reactor in an effort to maximize the utilization of the total energy output of the lamp and is similar to other applications of lamps in aqueous treatment systems. A maximum of 10 tubes were placed during each experiment. The reactor assembly was placed inside a photochemical reactor cabinet (92 cm H, 49 cm L, 56 cm W). The surfaces inside the cabinet were covered using aluminum foil to prevent light leakage.

2.3 Defluorination experiments

100 mL PFAS stock solutions were prepared as 2 mg L⁻¹ PFAS, 10 mM sulfite and pH 12 unless otherwise stated. The concentration of PFAS used in this study does not represent typical PFAS concentrations in natural waters but was selected to facilitate quantification by ion chromatography (IC). Equivalent mass concentrations were chosen because equivalent molar concentrations result in highly variable numbers of C–F bonds to be cleaved, and equivalent molar F concentrations results in variable number of non-fluorinated bonds. 10 min prior to irradiation, the lamp was ignited to allow for stabilization, as recommended by the manufacturer. Samples were subjected to irradiation for 0, 2, 5, 10, 15, 20 and 30 min. The temperature of the aqueous samples was monitored and recorded during the experiments using mercury immersion thermometer.

Four preliminary sets of experiments were conducted to determine the optimum photochemical conditions for PFAS defluorination. These experiments were conducted with PFOA and PFOS to assess the impact of initial sulfite, O₂, and MeOH concentrations and pH. First, samples were spiked with concentrations of sulfite from 1 to 20 mM at pH 12 and 2 mg L⁻¹ of PFOA or PFOS. The range was selected based on prior work by others which showed that 10 mM was an optimal sulfite dose with LP and HP lamps.^55,64 Sulfite concentration was monitored via IC, and samples were spiked with additional sulfite after 5, 10, 15 and 20 min of photolysis to maintain sulfite concentrations above 10% of the initial dose to ensure availability for the duration of the experiment. Second, after determining the optimum initial sulfite concentration, we varied pH between 2 and 12 with nitric acid or sodium hydroxide. Third, because MeOH is widely used to prepare PFAS solutions and because MeOH scavenges hydrated electrons,⁶⁵ MeOH was spiked at 0.05 to 20% to assess the potential to affect kinetics. Finally, we examined the impact of aqueous O₂ (1 to 8 mg O₂ per L) on PFAS defluorination rates because dissolved O₂ also scavenges e_aq⁻. Nitrogen (N₂) gas was bubbled through samples for 3 h to achieve O₂ concentrations below the detection limit. N₂ sparged water was also used for stock solutions used in these experiments.

The optimized conditions which were used in subsequent experiments were: 2 mg L⁻¹ PFAS, 10 mM initial sulfite dose, pH 12 ± 0.2. N₂ sparging was not conducted except for specifically identified experiments. With the optimized conditions, we conducted experiments which investigated defluorination kinetics and the effect of sample matrix (tap water, river water, and wastewater effluent) on defluorination rate over 30 min of UV/sulfite exposure. Tap water was collected in the laboratories of the University of Nevada, Reno and stored capped for 7 days at room temperature before use to ensure consumption of chlorine residual before experiments. A sample from the Truckee River in Reno, NV was collected in June 2020 and was immediately filtered with a polyethersulfone filter with nominal pore diameter of 0.45 μm. Wastewater effluent was collected from the filter effluent of a local wastewater reclamation facility in Reno, NV. The wastewater treatment plant consists of tertiary treatment including enhanced biological phosphorus removal, nitrifying trickling filters, and tertiary denitrification with added MeOH.

UV/sulfite experiments were conducted in triplicate. Dark and UV controls were also conducted simultaneously for each set of samples. Dark control sample tubes were wrapped with aluminum foil and placed next to photolysis tubes. No significant defluorination was observed in any dark controls (≤5%). UV sample tubes were exposed to irradiation but without added sulfite. Degradation in the UV controls was <5% (Table S3†). Thus, all results are reported without correction for controls.

2.4 Analytical methods

F⁻, sulfite, chloride, sulfate, nitrate and nitrite concentrations were analyzed using a Dionex ICS-3000 equipped with an anion exchange column (Dionex IonPac AG19 4 × 50 mm guard column and AS19 4 × 250 mm analytical column) and a conductivity detector (CD-20, Dionex). The mobile phase was 12 mM potassium hydroxide at 1 mL min⁻¹. The injection volume was 500 μL. The F⁻ detection limit was <5 μg L⁻¹. pH was measured with a pH meter (UB-7, Denver Instrument) calibrated daily. A Thermo Scientific Orion Star A2136 was employed to measure O₂. DOC was measured with a Shimadzu TOC-L (Shimadzu Corp., Kyoto, Japan) according to Standard Method 5310 B.⁶⁶

Pseudo first-order kinetics were from linear best fits of the natural log of F⁻ release vs. time. Data points prior to plateauing in F concentration were included in the fits. In all fits, the examined PFAS is “degraded” when all of the F present in the parent molecule is measurable as free F⁻ (i.e., 15 mM F⁻ indicates destruction of 1 mM PFOA). The calculated half-life, in this context, is the time required to cleave 50% of the C–F bonds present in the sample. This is a more rigorous definition of PFAS destruction compared to many other publications and likely results in the overestimation of required energy, in comparison, because PFAS are “destroyed” when measured by mass spectrometry once a single atom is cleaved, independent of the resulting product molecules/ions. We adopt this meaning of destruction (i.e., complete defluorination) because it was facilitated by the instrumentation (IC) but also because we believe that this is the definition that will be likely be used in future clean-up or regulatory actions due to the emerging bioactivity of shorter chain and lower molecular weight perfluorinated decomposition products.^62,67

To compare energy efficiency of UV/sulfite processes to published literature, we calculated the energy required (EE/O) to reduce the concentration of the targeted PFAS by one order of magnitude via complete defluorination (i.e., completely defluorinate 90% of the PFAS present, eqn (2) in kW h m⁻³).

	(2)

where P is the rated power (kW) of the MP UV lamp, t is the UV irradiation time (min), V is the volume of the sample in the reactor (15 mL) and C_i and C_f are initial and final concentration of the PFAS as F. Eqn (2) can be further simplified by substituting with base 10 logarithmic rate equation (k = −log(1 − C_f/C_i)/0.4343 × t) (eqn (3)).

	(3)

EE/O for each PFAS was estimated based on the irradiation received by the surface area of the quartz vial batch reactor (1.8 cm ID, 7 cm L). The reactor received 7% of total irradiation produced by the MP lamp based on the ratio of the vial's exposed surface area to a projected cylinder of irradiated surface 3 cm from the lamp.

3. Results and discussion

3.1 Impact of aqueous conditions on PFOA and PFOS defluorination

3.2 Defluorination kinetics of 22 PFAS by MP UV/sulfite

We proceeded to measure the defluorination kinetics of 22 PFAS using optimized sulfite, O₂, and H⁺ concentrations (Fig. 5). In Table 1, we summarize the pseudo first-order defluorination rate constants, half-lives, and linear regression coefficients for the 22 PFAS examined. R² values ranged from 0.82 to 0.99, indicating that the defluorination of these PFAS was well described by pseudo first-order kinetics. For the 11 PFCAs, defluorination rates ranged between 0.089 min⁻¹ for PFPrA (C₃) and 0.001 min⁻¹ for PFTeDA (C₁₄) resulting in half-lives of 7.8 to 578 min. The most poorly defluorinated PFCAs were PFDoA, PFTrDA and PFTeDA with observed rate constants of 0.007, 0.003 and 0.001 min⁻¹, respectively. Bentel et al. also examined degradation of PFAS under LP UV/sulfite and observed half-lives between 5.8 and 9.5 min for C₃ to C₈ PFCAs.⁵³ Half-lives calculated in this work indicate the time required to release 50% of F present in PFAS, whereas half-lives reported by Bentel et al. indicate the time required to degrade 50% of targeted PFAS (i.e., cause a change in m/z measured by mass spectrometry). Thus, it was expected that half-lives measured in this work would be greater than those reported by Bentel et al.

In general, PFCA defluorination rates decreased with increasing fluoroalkyl chain length. This behavior appeared to be exclusive to PFCAs and FTOHs, and the opposite was observed for PFSAs; defluorination rates increased with the increase in the fluoroalkyl chain length, although the number of included PFSAs in this groups was limited. This suggests separate mechanisms of e_aq⁻ attack and subsequent defluorination. PFCA degradation and defluorination mechanisms by e_aq⁻ attack have been investigated by Bentel et al.^53,64 At pH 12, decarboxylation and hydroxylation followed by HF elimination and hydrolysis dominate degradation of short-chain PFCAs. However, for longer-chain PFCAs, some transformation products react poorly with e_aq⁻ and thus reach a defluorination plateau. This is well reflected in our findings, where shorter chain PFCAs were degraded more rapidly than longer chain. Investigations of defluorination mechanisms of FTOHs and iodinated PFAS, which were poorly degraded in this work, have not been published and represent an opportunity for future research.

PFOS and PFHxS had shorter half-lives than PFCAs (22.1 and 44.1 min, respectively) but FTOHs, three iodinated PFAS, 6:2 FTO and 8:2 FTAC exhibited slow defluorination with half-lives ranging between 122 to 2311 min. 8:2 FTAC was the most recalcitrant PFAS in this study. Comparing C₈ analogues across functional groups results in the following order of decreasing defluorination rate constants: R–COOH (PFOA) > R–SO₃H (PFOS) > R–CH₂–CH₂–OH (FTOH) ≫ R–HCCH–I (TIFE) > R–HCCH₂ (6:2 FTO) > R–CH₂–CH₂–I (6:2 FTI). Overall, terminal functional groups played an important role in defluorination kinetics.

GenX is reported to be more toxic than PFOA, PFHxA, PFBA and PFOS and has been detected in various water surfaces around the globe.^73–75 GenX was the most rapidly defluorinated PFAS examined in this study (half-life of 4.3 min), and prior work has shown that the ether is particularly susceptible to e_aq⁻ attack.⁷⁶ These results are promising considering the increased use of GenX over legacy PFAS.

In Table 1, we report the EE/O to reduce the concentration of the targeted PFAS by one order of magnitude via complete defluorination (i.e., completely defluorinate 90% of the PFAS present), using the MP UV/sulfite system. EE/O in Table 1 are corrected to account for only 7% of the total irradiation energy produced by the lamp being incident to the vial. Table S5† contains the uncorrected EE/O. We also report both the total EE/O and the EE/O considering only the production of 254 nm light. The latter is provided for comparison to literature but should take into consideration EE/O in this work is calculated based on complete defluorination of the parent molecule. When all wavelengths were taken into account, including those which are not absorbed by sulfite (Table S4†), EE/O for C₂ to C₈ PFCAs ranged from 345 to 567 kW h m⁻³ and increased with increasing fluoroalkyl chain length. By comparison, Bentel et al.⁵³ found that EE/O for the same set of PFCAs irradiated with a monochromatic LP UV-sulfite system ranged from 20 to 457 kW h m⁻³ but was independent of chain length, likely due to the EE/O calculations being based on a single F loss, where those provided here are based on the more conservative complete defluorination. MP lamps are more energy intensive than LP and therefore their applications may be limited to specific concentrated effluents or situations where reactor residence time is a constraint.

3.3 Impact of environmentally relevant matrices on degradation kinetics

To assess the effect of water constituents typical of environmental samples we sampled tap water, river water, and wastewater effluent and conducted defluorination experiments with five PFAS. Table 2 contains the raw water quality characteristics of the water samples. To measure initial concentrations of PFAS/organofluorine which might interfere with spiked PFAS defluorination measurements, sample aliquots were exposed to UV/sulfite for 30 min without PFAS addition. Background organofluorine concentrations were ≤0.05 mg F per L, less than 4% of the spiked PFAS-F concentrations. Five PFAS were spiked at 2.0 mg L⁻¹ (PFPrA, PFPeA, PFOA, PFOS and GenX) and the defluorination results presented are corrected for the background F⁻ and organofluorine-F. The spiked PFAS concentration is significantly higher than would be expected in environmental samples but allows for comparisons to rate constants measured in clean samples.

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Tap water and river constituents had little or no impact on defluorination (<5% change from clean matrix after 30 min, Fig. 6). However, constituents contained in wastewater effluent reduced defluorination compared to the clean matrix.

Considering defluorination rate constants in the wastewater matrix (Fig. S3†), PFAS defluorination rates were reduced by a factor of 2.8, 2.0, 1.9, 1.7 and 3.5 times for PFPrA, PFPeA, PFOA, PFOS and GenX, respectively, compared to defluorination in a clean matrix. The inhibiting effects were more pronounced for the most rapidly degraded PFAS, with GenX being the most sensitive PFAS identified in this study. These results suggest that e_aq⁻ concentrations are reduced and/or scavenged by constituents present in wastewater effluent sample and the river and tap water samples had lower concentrations of potential scavengers (Table 2). The presence of DOC in wastewater effluent might been the major component responsible on defluorination reduction observed. DOC contains broad diversity of chromophores that are capable of absorbing light,⁷⁷ negatively impacting activation of sulfite by UV light and reducing the production of e_aq⁻. Moreover, several studies have indicated that dissolved organic matter sorbs PFAS, potentially reducing the interactions between e_aq⁻ and PFAS molecules.^54,78

4. Conclusions

MP UV/sulfite is rapid and effective in PFAS destruction and defluorination. Defluorination rates were found to be less sensitive to operational conditions (e.g., sulfite concentration, pH, MeOH, O₂) than has been observed in other published research utilizing LP lamps. The results presented here demonstrate that UV/sulfite systems may be feasible at scale if system input energy is high, potentially through arrays of LP lamps or implementation of MP lamps, because high energy input overcomes the challenges of pH adjustment and sparging to remove O₂. Defluorination rates varied between 0.0003 and 0.1604 min⁻¹ and were dependent on functional groups, fluoroalkyl chain length and the presence of additional e_aq⁻ scavengers such as DOC. PFCAs exhibited appreciable defluorination; rates increased with decreasing fluoroalkyl chain length. However, this was not observed for other PFAS. GenX was the most rapidly defluorinated PFAS examined, with a half-life of 4.3 min, and 8:2 FTAC was the most slowly degraded with half-life of 2311 min. Other PFAS such as FTOHs and iodinated PFAS were also slowly degraded. If applied to water treatment, attention to dissolved water constituents should be considered due to inhibition. This research demonstrated the potential for rapid and robust defluorination of PFAS, which may outweigh the energy cost in specific applications where short residence times are required.

Conflicts of interest

There are no conflicts to declare.

Acknowledgements

This research was partially supported by the Strategic Environmental Research and Development Program under Grant ER19-1214. Views, opinions, and/or findings contained in this report are those of the authors and should not be construed as an official Department of Defense position or decision unless so designated by other official documentation.

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Footnote

† Electronic supplementary information (ESI) available. See DOI: 10.1039/d1ew00221j

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