Complete Defluorination and Mineralization of Perfluorooctanoic Acid

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Cite This: Environ. Sci. Technol. 2019, 53, 8302−8313

Complete Defluorination and Mineralization of Perfluorooctanoic Acid by a Mechanochemical Method Using Alumina and Persulfate Nan Wang,†,§ Hanqing Lv,†,§ Yuqi Zhou,† Lihua Zhu,*,† Yue Hu,‡ Tetsuro Majima,† and Heqing Tang*,‡ †

College of Chemistry and Chemical Engineering, Huazhong University of Science & Technology, Wuhan 430074, P. R. China College of Resourcesand Environmental, South-Central University for Nationalities, Wuhan 430074, P. R. China

‡

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S Supporting Information *

ABSTRACT: Perfluorooctanoic acid (PFOA) is a persistent organic pollutant that has received concerns worldwide due to its extreme resistance to conventional degradation. A mechanochemical (MC) method was developed for complete degradation of PFOA by using alumina (Al2O3) and potassium persulfate (PS) as comilling agents. After ball milling for 2 h, the MC treatment using Al2O3 or PS caused conversion of PFOA to either 1-H-1-perfluoroheptene or dimers with a defluorination efficiency lower than 20%, but that using both Al2O3 and PS caused degradation of PFOA with a defluorination of 100% and a mineralization of 98%. This method also caused complete defluorination of other C3∼C6 homologues of PFOA. The complete defluorination of PFOA attributes to Al2O3 and PS led to the weakening of the C−F bond in PFOA and the generation of hydroxyl radical (•OH), respectively. During the MC degradation, Al2O3 strongly anchors PFOA through COO−−Al coordination and in situ formed from Lewis-base interaction and PS through hydrogen bond. Meanwhile, mechanical effects induce the homolytic cleavage of PS to produce SO4•−, which reacts with OH group of Al2O3 to generate •OH. The degradation of PFOA is initiated by decarboxylation as a result of weakened C−COO− due to Al3+ coordination. The subsequent addition of •OH, elimination of HF, and reaction with water induce the stepwise removal of all carboxyl groups and F atoms as CO2 and F−, respectively. Thus, complete defluorination and mineralization are achieved.

■

INTRODUCTION

Mechanochemical (MC) degradation is a promising method for the disposal of organic pollutants in solids. In an MC treatment, the degradation reaction usually begins with the MC activation of milling agents by particle crushing, buildup of structural defects, rupture of bonds, heating, etc., under intensive mechanical stresses. The efficiency of the MC degradation of organic pollutants is improved by using appropriate additives. For example, CaO,14 CaO-SiO2,15 Fe,16 zerovalent metal (e.g., Fe, Al)-SiO2,17 persulfate (PS),18 and PS−CaO19 have been used as efficient additives in MC degradation of chlorinated and brominated organic pollutants, although these additives are not efficient enough individually for the MC degradation of PFOA (SI Table S1). Yan et al. investigated the MC degradation of a chlorinated polyfluorinated ether sulfonate (F-53B, a PFOS alternative) by using sodium persulfate and NaOH as comilling agents, and found that the defluorination efficiency of F53−B was 54% after ball milling for 8 h.20 Zhang et al. reported that ball milling for 8 h produced defluorination efficiency of PFOA and F-53B greater

Perfluorooctanoic acid (C7F15COOH, PFOA), perfluorooctanesulfonate (C7F15SO3H, PFOS), and other fluorosurfactants have been widely applied in textiles, paper and packaging materials, cookware, and fire-fighting foam productions. Due to their disposal, these fluorocarbons are frequently detected in aquatic environments,1 soils, sludge, and sediments.2 Because PFOA and PFOS are persistent organic pollutants (POPs),3 their elimination is necessary. PFOA is thermally and chemically stable because of high dissociation energy of C−F bonds (533 kJmol−1).4 To induce the degradation of PFOA in aqueous media, several chemical approaches have been explored, such as UV-sulfite reduction,5,6 photo- and electro-Fenton oxidation,7,8 photocatalytic oxidation,9 and sulfate radical anion (SO4•−)-based oxidation.10−12 However, few methods have been developed for the degradation of PFOA in solids. Incineration at temperatures above 500 °C is a practical method to treat organic pollutants in solid, although the thermal treatment of PFOA produces undesirable greenhouse gases and corrosive gases such as CF4, C2F4, and HF.13 Thus, development of complete and environmentally friendly disposal technologies are urgently needed for treating PFOA in solids. © 2019 American Chemical Society

Received: Revised: Accepted: Published: 8302

January 23, 2019 May 29, 2019 May 31, 2019 May 31, 2019 DOI: 10.1021/acs.est.9b00486 Environ. Sci. Technol. 2019, 53, 8302−8313

Article

Environmental Science & Technology

10 mm (the total mass of balls was 210 g). When PS and Al2O3 were used as comilling agents, the molar ratio of Al2O3/PS (nAl2O3/nPS) was set at 5 unless otherwise stated. The mass of the total milled sample was 4.2 g, and the ball-to-sample mass ratio was fixed at 50:1. After the milling materials were well predistributed in a pot, the pot was sealed tightly and the planetary ball mill was operated at 350 rpm under atmospheric conditions, with automatic change in the direction of rotation every 15 min. At given time intervals, the ball milling was stopped and solid samples were taken out for chemical analysis, unless specified elsewhere. Replicate runs (n ⩾ 3) were carried out for each test, and the relative standard deviations were less than 5%. For isotopic labeling experiments, both the PS and Al2O3 were preheated at 100 °C for 24 h and kept in a desiccator prior to MC degradation experiments. H218O (1 mL) was mixed with Al2O3 (10 g) in a desiccator for 1 h to ensure full oxygen exchange between H218O and surface OH on the Al2O3, followed by vacuum drying at 60 °C for 1 h. Then, 18Olabeled Al2O3 powders were used for the MC treatment of PFOA. Analysis. Each milled sample (0.020 g) was extracted with ethanol (three times, for a total of 10 mL) under ultrasound irradiation (KQ-200 KDE, Kunshan Ultrasonic Instruments, China) for 10 min. After removing the solids by centrifugation, the collected solution was filtered through a 0.22 μm membrane and then subjected to HPLC analysis. Similar to our previous work,27 the PFOA concentration was measured with HPLC on an Ultimate 3000 series system (Dionex, Idstein, Germany) equipped with a Corona Ultra RS charged aerosol detector (CAD, Thermo Scientific, Bellefonte, PA). The analysis was performed on a Symmetry C8 column (150 mm × 3.9 mm, 5 μm particle size; Waters, Ireland). Column oven temperature was set at 30 °C. The mobile phase was 30% ammonium acetate (5 mmol L−1 with pH adjusted to 5.0 ± 0.2 by acetic acid) and 70% methanol by volume, the flow rate was 1 mL min−1, and the injection volume was 40 μL. The degradation efficiency of PFOA (DGEPFOA) is calculated from eq 1,

than 80% by adding KOH as a milling agent with a KOH/ PFOA molar ratio of 170:1, but less than 20% by adding NaOH.21,22 The relatively poor defluorination in the MC degradation of PFOA is closely related to the chemical inertness of C−F bonds. Activation of a C−F bond and the following cleavage involve the formation of thermodynamically favorable bonds such as H−F, B−F, Al−F, Si−F, and Ge−F bonds.23,24 Recently, we have found that the mechanocaloric effect promotes the dehydration of the Al2O3 surface to produce coordinatively unsaturated Lewis acid sites, inducing the activation of C−F bonds to convert PFOA to a valuable product, 1-H-1-perfluoroheptene.25 Cagnetta et al. reported that the use of La2O 3 as a milling agent converted perfluorinated compounds into LaOF, a luminescent material.26 This MC method is preferable for the degradation of solid PFOA, because not only is PFOA degraded, but organofluorine blocks in PFOA are also recovered for synthesizing polyfluoroalkenes or LaOF. However, MC treatment is not appropriate for degradation of PFOA at ratios lower than 20% in solids because the economy of this method becomes poor for recovery of the fluorine resource. In most cases, the PFOA content is too low to be recovered as a resource. Therefore, a degradation of PFOA with complete defluorination in solids to harmless substances is desirable. We recently have reported using the MC method with PS to cause the homolytic cleavage of the peroxide bond in PS to produce SO4•−, inducing the decomposition of brominated organic pollutants.18 Considering the activation effect of Al2O3 on C−F bonds of PFOA and generation of SO4•− from the dissociation of PS by the MC treatment, we anticipated that Al2O3 and PS might lead to complete MC degradation and defluorination of PFOA. In the present work, we developed an MC method for degradation of PFOA by using both Al2O3 and PS as comilling agents, and clarified the reaction mechanism.

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EXPERIMENTAL SECTION Chemicals and Reagents. Perfluorooctanoic acid (PFOA, C7F15COOH, 98% in purity) was purchased from J&K China Chemical, Beijing, China. Shorter chain perfluorocarboxylic acids (PFCAs) including perfluoroheptanoic acid (PFHpA, >98%), perfluorohexanoic acid (PFHxA, >98%), and perfluoropentanoic acid (PFPA,>98%) were purchased from Tokyo Chemical Industry, Shanghai, China. Alumina (Al2O3, 99.9%), 5,5-dimethyl-1-pyrroline (DMPO, analytical grade), and diphenyl-1-picrylhydrazyl (DPPH•, 97%) was obtained from Aladdin, Shanghai, China. H218O with an isotope abundance of 98% was purchased from Jiangsu Huayi Isotopes Chemical Limited. Potassium persulfate (PS, analytical grade) was obtained from Sinopharm Chemical Reagent Co., Ltd., Shanghai, China. All solvents used for extraction and highperformance liquid chromatography (HPLC) analysis were analytical grade and HPLC grade, respectively. Milli-Q water with conductivity of 18.2 MΩ cm (Merck Millipore GmbH, Germany) was used in the present work. Degradation Experiments. The MC degradation of PFOA was conducted in a planetary ball mill at room temperature (BM4, Beijing Grinder Instrument, China) by using Al2O3 and/or PS as milling agents. Typically, 0.25 g PFOA and 3.95 g milling agents (PS, Al2O3, and/or their mixtures) were mixed in the stainless steel pot (250 mL) of the ball milling machine, which was then filled with 18 stainless steel balls with a diameter (d) of 6 mm and 10 balls with d =

ij nPFOA, t yzz zz × 100% DGE PFOA = jjjj1 − z n PFOA,0 k {

(1)

where nPFOA,0 and nPFOA, t are moles of PFOA at reaction times 0 and t, respectively. In addition, the ethanol-extracted solutions were analyzed using an 1100 LC/MSD trap system (Agilent, Santa Clara, CA) with an electrospray ionization (ESI) source to identify possible polar intermediates. To identify volatilized organic intermediates, both products in the reactor and gaseous products in the head space of the reactor were collected by using acetone as a solvent. The extract was filtered through a 0.22 μm membrane and then directly analyzed by GC/MS (Thermo Fisher, U.S.) under the same conditions reported in our previous work.25 For the analysis of the CO2 produced, the gases were collected in 20 mL NaOH (0.5 mol L−1) after the ball milling. The CO2 absorbed in NaOH was converted into CO32− and then detected by ion chromatography (IC) on a Dionex ICS-1500 equipped with a CD 25 conductivity detector, IonPac AS 23 column and KOH as the eluent (28 mmol L−1, 1.0 mL min−1). In the isotopic labeling experiments, the produced CO2 was analyzed using Fourier transform infrared (FT−IR) spectrometry (Bruker, Germany) by introducing the gaseous products 8303

DOI: 10.1021/acs.est.9b00486 Environ. Sci. Technol. 2019, 53, 8302−8313

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Environmental Science & Technology

Figure 1. Time profiles of degradation (a) and defluorination (b) of PFOA (0.1 g) in kaolin soil (1.5 g) under 350 rpm and a ball-to-sample mass ratio (mb/ms) of 50 in the presence of Al2O3 (5.7 g), PS (5.7 g), or the mixture of Al2O3 (2.4 g) and PS (3.3 g). All milling agents were added in one time before milling, except for the system of Al2O3+PS (1/3 × 3), in which PS and Al2O3 were divided into three equal portions and added at reaction time of 0, 30, and 60 min.

ments, China) for 10 min. After centrifugation and filtration through a 0.22 μm membrane, 1.5 mL filtrate was mixed with 1.0 mL phosphate buffer (pH 6.86) and 0.5 mL KI (2 M). After 15 min, 0.20 mL of this mixture was diluted to 10 mL with water and then measured at 355 nm on a UV−vis spectrophotometer (Cary 50, Varian,Palo Alto, CA). The recovery yields of PS were greater than 98% with the standard addition method by spiking PS (1.37 g) into a mixture of PFOA (0.25 g) and Al2O3 (2.58 g). The utilization efficiency of PS (ηPS) is defined as the ratio of the stoichiometric requirement to the actual consumption, which is calculated from the moles of generated F− and consumed PS (ndetected F−, t and nPS,0 − ndetected PS, t, respectively) (eq 3),

into the gas cell. The distribution of 18O in CO2 was analyzed not only by GC/MS (Agilent 7890B-5977A GC/MSD) with a DB-5MS column (30 m × 0.25 mm; 0.25 μm particle size) and quadruple mass analyzer, but also by isotope ratio mass spectrometry after converting the gaseous CO2 to a precipitate of MnCO3. After the MC reaction, the generated CO2 was absorbed by a NaOH solution, followed by isolating the milled powders using centrifugation and filtration. The filtrate, with a pH approximately 7.5, was adjusted to pH 11 with NaOH (0.5 mol L−1), followed by adding MnSO4·H2O (0.08 g) to yield a MnCO3 precipitate. The 18O isotope abundance in MnCO3 was analyzed by using an elemental analyzer (EA IsoLink CN/ OH, Thermo Fisher USA) coupled to an isotope ratio mass spectrometer (Delta V Advantage, Thermo Fisher). In the reaction furnace of the elemental analyzer at 1380 °C, the evolved oxygen in MnCO3 reacted with carbon to yield CO, which was separated and then transported into an isotope ratio mass spectrometer. Because the conversion of MnCO3 to CO strongly depends on temperature, two reference compounds including CO and MnCO3 that obtained from the reaction of Na2CO3 with MnSO4 (named ref-MnCO3) were subjected to the same measurement conditions. To determine the amount of F− released from PFOA, the milled samples (20 mg) were dispersed in 10 mL water and 0.5 mL aqueous ammonia. The suspensions were centrifuged and filtered through a 0.22 μm membrane. The filtrate then was diluted to 50 mL with 10 mL total ionic strength adjustment buffer (TISAB, consisting of HAc-NaAc and 10 g L−1 sodium citrate) and water. Finally, F− was measured with a F− selective electrode. The recovery yields of F− were greater than 95% as determined using the standard addition method by spiking NaF (0.3 g) into 3.95 g Al2O3 or a mixture of PFOA (0.25 g), K2SO4 (1.57 g), and Al2O3 (2.58 g) (SI Table S2). The defluorination efficiency of PFOA (DFEPFOA) is evaluated from eq 2, DFE PFOA =

ndetectedF−, t nF,0

× 100% =

ndetectedF−, t nPFOA,0 × 15

ηPS =

αndetectedF−, t nPS,0 − ndetectedPS, t

× 100% (3)

where ndetected PS, t and nPS,0are moles of PS detected experimentally and calculated from the initial PS, respectively, and α is the stoichiometric ratio of the moles of generated F− to the moles of consumed PS. When necessary, solid samples were taken at specified time intervals during the ball milling and then were characterized using spectroscopic techniques. FT−IR spectra were measured on an Equinox 55 (Bruker, Germany) with the KBr disk method from 400 to 4000 cm−1. X-ray photoelectron spectrometry (XPS) was conducted on Kratos AXIS-ULTRA DLD-600W (Shimadzu, Japan), in which the binding energy is calibrated using the C 1s peak as 284.6 eV. 19F and 27Al MAS NMR experiments were performed at room temperature on a Bruker Avance III 500 MHz spectrometer (Bruker, Germany) equipped with a 2.5 mm Bruker 1H/19F/X probe. The samples were loaded into a zirconia rotor and spun at an MAS rate of 20 kHz. A Bruker EMX-nano electron spin resonance (ESR) instrument was employed to detect radicals with DMPO as the spin trapping reagent. Prior to measurements, a small amount of DMPO instead of PFOA was added to the milling mixture. After being milled for 15 min at 300 rpm, a portion of the milled sample (ca. 0.1 g) was directly collected and another portion (20 mg) of the milled mixture was dispersed in 10 mL water. Both the solid and water-extracted samples were measured under the following conditions: center field, 3480 G; sweep width, 200 G; microwave frequency, 9.77 GHz; modulation frequency, 40 kHz; power, 10 mW; scan times, 3.

× 100% (2) −

where ndetected F−, t and nF,0 are moles of F detected experimentally and F calculated from the initial PFOA (mol), respectively. The residual PS in the milled samples was determined with a KI spectrophotometric method.28 Prior to the measurement, the milled samples (20 mg) were dispersed in 10 mL water under ultrasound (KQ-200 KDE, Kunshan Ultrasonic Instru8304

DOI: 10.1021/acs.est.9b00486 Environ. Sci. Technol. 2019, 53, 8302−8313

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Environmental Science & Technology

Figure 2. Time profiles of degradation (a) and defluorination (b) of PFOA (0.25 g) under 350 rpm and mb/ms of 50 with milling agents such as PS (3.95 g), Al2O3 (3.95 g), or the mixture of Al2O3 (2.58 g) and PS (1.37 g) at nAl2O3/nPS of 5. Effects of nAl2O3/nPS (c) and mb/ms (d) on degradation (triangles) and defluorination ratios (circles) of PFOA in MC−Al2O3−PS at a 2 h milling. (e) Time profiles of the temperature on the surface of milled samples during the milling. (f) Time profiles of defluorination efficiency of PFOA with PS or the mixture of PS and Al2O3 in solid state by heating to 100 °C.

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RESULTS AND DISCUSSION

degradation experiments were carried out using pure PFOA as a model pollutant. Figure 2a and b compare the MC degradation and defluorination of powdered PFOA (0.25 g) by using 3.95 g milling agents. After a 2 h milling with either PS (nPS/nPFOA= 24) or Al2O3 (nAl2O3/nPFOA = 64), the degradation efficiency of PFOA was greater than 60%, but the defluorination extent was lower than 20%. When using comilling agents at nAl2O3/nPS = 5, both degradation and defluorination efficiencies reached 100% after 2 h milling. Moreover, at any time, the defluorination extent of PFOA was close to the degradation in MC−Al2O3− PS, thus confirming that successive degradation and defluorination continued until there is a complete removal of F atoms. In addition, the fate of fluorine was identified by solid-state nuclear magnetic resonance (SS NMR) and XPS measurements. High-resolution XPS spectra of F 1s and NMR spectra of 19F and 27Al demonstrated that F− eliminated from PFOA in

Complete Degradation and Defluorination of PFOA using Mechanochemistry. Figure 1a and b illustrate the MC degradation and defluorination of PFOA (0.1 g) in kaolin soil (1.5 g) with PS, Al2O3 or a mixture of PS and Al2O3 as milling agents (5.7 g). After milling for 2 h, all the added PFOA in the three cases was nearly degraded, but defluorination efficiencies calculated from the generated F− were 10%, 27% and 85% in MC−Al2O3, MC−PS, and MC−Al2O3−PS, respectively. When Al2O3 (2.4 g) and PS (3.3 g) were divided into three equal portions and added for the three reaction times of 0, 30, and 60 min, the degradation and defluorination extents of PFOA achieved 99% and 96% at 2 h, respectively. This indicates that the use of both Al2O3 and PS favors the MC defluorination of PFOA. Because the soil matrix will bring more difficulties in clarifying the roles of Al2O3 and PS, the subsequent 8305

DOI: 10.1021/acs.est.9b00486 Environ. Sci. Technol. 2019, 53, 8302−8313

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Environmental Science & Technology

Figure 3. Distributions and balance of carbon during MC degradation of PFOA in (a) MC−Al2O3, (b) MC−PS, and (c) MC−Al2O3−PS. (d) Removal efficiency of total organic carbon (TOC) during MC degradation of PFOA. The measured TOC in (b) includes that in the generated organic products (such as fluoroketones, fluoroethers, and fluoroesters) and residual PFOA.

MC−Al2O3−PS is bonded to the Al2O3 surface to form Al−F bond (SI Text S1, Figures S1 and S2). This provides additional merit for reducing the secondary pollution because the bonding of F− on Al2O3 reduces the dissociation of F− into water and thus avoids the possible promotion effect of bone disease from excess F−. Figure 2c shows the effect of nAl2O3/nPS on the degradation and defluorination ratios of PFOA after a 2 h MC treatment. As nAl2O3/nPS increases from 0 to 2, the degradation extent of PFOA increases from 61% to 96%. A further increase in nAl2O3/ nPS to 5 results in a degradation ratio for PFOA of 100%. However, a different volcano plot is observed for the effect of nAl2O3/nPS on the defluorination efficiency of PFOA. The defluorination ratio of PFOA increases from 6.2% to 98% when nAl2O3/nPS increases from 0 to 3 and reaches 100% at nAl2O3/ nPS= 5. After that, the defluorination ratio of PFOA significantly decreases to 42% as nAl2O3/nPS increases to 10. The use of Al2O3 alone leads to a defluorination ratio of 18%, even though all the added PFOA is removed. The above observations are related to different roles of comilling agents on the degradation and defluorination of PFOA. Al2O3 exhibits better performance than PS on the degradation of PFOA via a decarboxylation process as a result of weakened C−COO− due to Al3+ coordination (Figure 2a), leading to that a larger nAl2O3/ nPS caused a higher degradation ratio of PFOA. However, neither the in situ generated electrons in MC−Al2O3 nor the generated SO4•− in MC−PS was efficient in the defluorination of PFOA (Figure 2b). The complete defluorination of PFOA is attributed to Al2O3 and PS, which caused the weakening of the

C−F bond in PFOA and the generation of hydroxyl radicals (•OH), respectively. During ball milling, mechanical effects induce the decomposition of PS into SO4•−, which reacts with surface OH− (OH−surf) on Al2O3 to generate •OHsurf. As nAl2O3/nPS decreases, the total amount of OH−surf on Al2O3 decreases. In contrast, as nAl2O3/nPS increases, the amount of PS decreases. Either of them would result in the insufficient generation of •OHsurf from the reaction of SO4•− with OH−surf. Thus, a complete defluorination of PFOA requires a moderate value of nAl2O3/nPS. Figure 2d illustrates the effect of the ball-to-sample mass ratio (mb/ms) on the MC degradation of PFOA at a fixed nAl2O3/nPS/nPFOA of 42:8.4:1. As mb/ms increases from 15 to 50, both the degradation and defluorination ratios of PFOA at 2 h increases from approximaterly 35% to 100%. However, slight increases in the degradation and defluorination ratios of PFOA are observed with an increase in mb/ms to 75 or more. Larger mb/ms values can provide higher mechanical energy or a stronger mechanocaloric effect, which may induce an increase in temperature. However, as mb/ms varied from 25 to 100, the change in temperature on the surface of the milled samples was almost the same and the maximum temperature was below 50 °C (Figure 2e), due to the instantaneous temperature produced by ball milling having a short lifetime of 2−10 ns.29 After the ball milling, both PS and Al2O3 were not molten (SI Figure S3), because the surrounding temperature was much lower than their melting points. In a control test, when the powdery mixtures of PFOA, Al2O3 and PS were heated at 100 °C for 2 h, the defluorination efficiency of PFOA was negligible (Figure 2f). Thus, the activation of solid-to-solid 8306

DOI: 10.1021/acs.est.9b00486 Environ. Sci. Technol. 2019, 53, 8302−8313

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Environmental Science & Technology

Al2O3, MC−PS, and MC−PS−Al2O3, respectively (Figures 3a∼c). Throughout the MC treatment, the total carbon included in the residual PFOA, organic intermediates and produced CO2 was almost equal to the initial value for the three tested reactions. Thus, it is emphasized that PFOA was converted to either CF3(CF2)4CFCFH in MC−Al2O3, or dimeric molecules in MC−PS, whereas the degradation to CO2 was dominant in MC−PS−Al2O3. Moreover, the TOC removal efficiency of PFOA in MC−PS−Al2O3 was 98%, being 5.4 and 8.9 times higher than those of MC−Al2O3 (18%) and MC−PS (11%), respectively. Interaction of PFOA with Comilling Agents and the Transformations. After mixing PFOA with PS and/or Al2O3 by manually grinding in an agate mortar, a small portion of the sample was compressed into a KBr pellet for the FT−IR measurement. Table 1 summarizes the dominant absorption

reactions by ball milling is not a simple heat activation mechanism. The high energy impact between surfaces of the balls not only produces a transient high temperature, but also initiates the rotation, alignment, and reorientation of molecules, all of which facilitate changes in the solid surface and chemical bonds. A larger mb/ms can increase the total number of impacts between balls to achieve reaction completion for a given milling time, and any further increase of mb/ms can only act to accelerate to accelerate reaction rate.30,31 Mineralization of Degradation Intermediates. The degradation intermediates produced during the 1 h MC degradation of PFOA were extracted with ethanol or acetone as described in the Experimental Section. The organic intermediates in the ethanol-extracted solutions were identified by LC/MS, and those in the acetone-extracted solutions were identified by GC/MS. When Al2O3 was used as a milling agent, three products were detected in the acetone extracts from milled samples, where CF3(CF2)4CFCFH was observed as a major product in a yield of 87% with the almost constant selectivity at 88% after a 2 h milling (Figures 3a and SI Figures S4a, and S5). This indicates that the transformation to CF3(CF2)4CFCFH was the main degradation pathway of PFOA in MC−Al2O3. When using PS as a milling agent, large amounts of fluoroketones, fluoroethers, and fluoroesters in the acetoneextracted solutions was observed as well as trace amounts of shorter chain PFCAs (such as perfluorohexanoic acid (C 5 F 11 COOH, PFHxA) and perfluoroheptanoic acid (C6F13COOH, PFHpA) in ethanol-extracted solutions (SI Text S1, Figures S4b, S6, and S7). Due to the lack of authentic compounds, we did not quantify the individual intermediates, but measured the total organic carbon (TOC) in the milled sample. As shown in Figure 3b, the TOC resulting from the generated organic products and residual PFOA was 89%, although 61% of the added PFOA was degraded after a 2 h milling. When both Al2O3 and PS were used as comilling agents, two components were detected in the ethanol-extracted solutions (SI Figure S8). The first component was detected at 3.9 min with m/z = 314.9(C5F11COO−), and 272.1(C5F11−) was identified as PFHxA (C 5 F 11 COOH), and the second component was detected at 5.6 min with m/z = 362.8(C6F13COO−) and 319.7(C6F13−) was identified as PFHpA (C6F13COOH). Four main polyfluorocarbons (PFCs) were detected in the acetone-extracted solutions (SI Figure S4a). The component that was detected at 1.40 min with the mass fragments of 219.12 (CF3(CF2)3), 169.17 (CF3(CF2)2), 119.15 (CF 3 CF 2 ) and 69.2 (CF 3 ) was assigned to CF3(CF2)2CF2H (SI Figure S9a). Similarly, the other three components were identified as CF3(CF2)3CFCF2 at 1.46 min, CF3(CF2)4CFCFH at 1.50 min, and CF3(CF2)5CF2H at 1.60 min, respectively (SI Figures S9b−d). The maximum ratio of total organic intermediates was less than 10% of the initial PFOA level, and their accumulation gradually decreased to zero as the ball milling time was prolonged to 2 h (Figure 3c). These comparisons confirm that using both Al2O3 and PS not only alters the pathway of PFOA degradation but also significantly reduces the accumulation of intermediates. In addition, CO2 yields were measured in all three tested mixtures when the gaseous products were collected in NaOH solution. After a 2 h milling, the ratios of CO2 to the initial amount of carbon in PFOA were 18, 7.9, and 98% for MC−

Table 1. Characteristic FT−IR Peaks of PFOA, PS, Al2O3, and Their Mixtures before Milling absorption peaka /cm−1 samples PFOA

ν(C = O/ COO−) 1768

ν(CF)

PFOA−Al2O3

1768, 1722(w) 1676, 1413 (w)

1236, 1204, 1147 1243, 1211, 1151

PS−Al2O3 PFOA−Al2O3− PS a

1681, 1415 (w)

νs(SO3−)

1303, 1276

1108

1303, 1276

1108(w)

1303, 1276 1303, 1276

1108(w)

1236, 1204, 1147

PS Al2O3 PFOA−PS

νas(SO3−)

1243, 1211, 1151

1108(w)

The letter “w” in parentheses represents the weak absorption peak.

peaks together with their assignments. In FT−IR spectrum of PFOA−PS, a new and weak peak appeared at 1723 cm−1 in addition to the strong absorption peak for the carboxyl group of PFOA (Table 1, SI Figure S10a). Meanwhile, the symmetric stretching (νs) peak of SO3− at 1108 cm−1 significantly decreased in intensity. This indicates that most PFOA molecules exist in the form of carboxylic acid, while a small amount of PFOA is bound to PS through the intermolecular hydrogen bond between the SO3− of PS and the COOH of PFOA.32 For PFOA−Al2O3, the stretching peak of CO (1768 cm−1) in PFOA disappeared, and νs and the asymmetric stretching peak (νas) of COO− appeared at 1413 and 1676 cm−1, respectively, with Δν (=νas−νs) of 263 cm−1. The stretching peaks of the C−F units moved to higher wavenumbers by 4∼7 cm−1 (Table 1, SI Figure S10b). These changes are related to both the COO−···Al coordination in a monodentate mode through the neutralization between COOH and surface OH on Al2O3 (>AlOH) (eq 4) and hydrogen bond between the F of PFOA and > AlOH (eq 5).25,33,34 After mixing the three components, the spectral changes related to the carboxyl group and C−F units were similar to those in PFOA−Al2O3 but different from those in PFOA−PS (Table 1, SI Figure S10c). This indicates that PFOA is bound to the Al2O3 surface rather than that of PS. The deprotonation of PFOA causes the dissociation of hydrogen bonding between PFOA and PS, leading to the recovery of the νs peak of SO3− at 1108 cm−1. Unlike the 8307

DOI: 10.1021/acs.est.9b00486 Environ. Sci. Technol. 2019, 53, 8302−8313

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Environmental Science & Technology

Figure 4. High resolution XPS spectra of (a) Al 2p and S 2p for PFOA−Al2O3−PS before and after the ball milling for 15 and 30 min.

Figure 5. Normalized UV−vis absorption spectra of DMSO-extracted solutions from the milled samples in the presence of DPPH• (a), where the absorbances of the peaks in the region of 200∼600 nm are normalized to 1.0. ESR spectra of milled samples in the presence of DMPO (b).

expectation, the weak peak at 1108 cm−1 in PFOA−Al2O3−PS was observed to indicate the presence of other hydrogen bond donors. After mixing PS and Al2O3, the νs peak of SO3− at 1108 cm−1 weakened (Table 1, SI Figure S10d), implying that the surface OH of Al2O3 and the SO3− of PS act as hydrogen bond donors and acceptors, respectively (eq 6). Thus, it is concluded that Al2O3 anchors the PFOA and PS in ternary mixtures and the two interaction modes including PFOA− Al2O3 and Al2O3−PS do not influence each other. The characteristic absorption peaks of PFOA and PS in MC− Al2O3−PS gradually decreased as the ball milling proceeded. Moreover, the rates of decrease were faster than those for MC−PS (SI Figure S11), which suggests that the coexistence of Al2O3 promotes the conversion of PFOA and PS. > AlOH + C7F15COOH → > Al3 +···− OOCC7F15 + H 2O

eliminated from the PFOA and then bonded to the Al2O3 surface. The sample before milling showed an S 2p peak at 170.3 eV, which was assigned to S2O82− (Figure 4b).37 After ball milling for 1 h, the S 2p peak gradually moved to 169.3 eV to be in the proximity of the binding energy of inorganic sulfates.38 This supports the hypothesis the conversion of S2O82− to SO42− took place during the MC degradation of PFOA. Identification of Reactive Species. During the ball milling, PS is activated to produce SO4•−,18 and the metal oxide generates oxygen vacancies and free electrons due to a collateral effect of losing lattice oxygen.25,39,40 To confirm such reactive species during the milling of Al2O3 and PS, we first employed DPPH• as a probe of free electrons and radicals. When the mixture of DPPH• and Al2O3 is subjected to the ball milling, the absorption peak of DPPH• at 525 nm significantly decreased with the appearance of a new peak approximately 425 nm corresponding to DPPH− (Figure 5a),25,39 which indicates the generation of free electrons on the milled Al2O3 (eq 7). However, no DPPH− was observed by using PS and Al2O3 as comilling agents, which suggests that the adsorbed PS as a strong oxidant effectively traps the electrons generated on Al2O3 surface (eq 8). In addition, MC−Al2O3−PS showed a broad absorption peak approximately 440 nm, which is different from that of MC−PS (Figure 5a), but similar to that of the coupling product between DPPH• with •OH.41

(4)

> AlOH + > Al3 +···− OOCC7F15 → > Al3 +···− OOCC7F15 ··· HOAl
AlOH + −OS(O2 )OOSO−3 → > AlOH ··· −OS(O2 )OOSO−3

XPS measurements were conducted to check the elemental changes of Al and S in MC−Al2O3−PS. Before the ball milling, the Al 2p peak approximaterly 74.8 eV was deconvoluted into two peaks at 73.9 and 75.0 eV corresponding to Al−O and Al− OH in Al2O3, respectively (Figure 4a).35 After ball milling for 15 min, the Al−O peak gradually decreased in intensity and a new peak appeared at 76.4 eV, which was assigned to F− bonded to Al (F−Al).36 With an increase in the milling time to 60 min, the Al−O peak became weaker and the intensity of F− Al peak increased. This indicated that more F atoms were

milling

8308

> Al − O ⎯⎯⎯⎯⎯⎯→ > Al − VO(2e−) + 0.5O2

(7)

e− + S2O82 −surf → SO4•−surf + SO4 2 −

(8)

DOI: 10.1021/acs.est.9b00486 Environ. Sci. Technol. 2019, 53, 8302−8313

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Environmental Science & Technology

Figure 6. (a) Time profiles of defluorination efficiency of PFOA in MC−Al2O3−PS and MC−PFOA−(18OH−Al2O3)−PS. (b) FT−IR and (c) mass spectra of collected gaseous products from MC degradation of PFOA. (d) Isotopic oxygen composition (δ18O) of MnCO3 obtained from degradation products of PFOA in MC−Al2O3−PS at a 2 h milling (S1-MnCO3) and MC−(18OH−Al2O3)−PS at a 4 h milling (S2-MnCO3). CO2 obtained from the reaction Na2CO3 with HNO3 and indoor air, commercial CO gas and MnCO3 obtained from the reaction Na2CO3 with MnSO4 (ref-MnCO3) were given as references in panels (c) and (d). milling

The radical intermediates were also investigated by ESR spectroscopic measurement of solids recovered from the milled samples in the presence of DMPO as a spin trapping reagent. As shown in Figure 5b, MC−Al2O3 showed no detectable signal in either solid or solution samples. The milled sample of PS (1.0 g) in the presence of DMPO exhibited the strong characteristic peaks of a DMPO−SO4•− adduct with hyperfine splitting constants of aN = 13.2 G, aH= 9.6 and 1.48G (Figure 5b), which confirms the decomposition of PS into SO4•− under the ball milling conditions (eq 9).42,43 For the mixture of Al2O3 (0.7 g) and PS (0.3 g), the DMPO−SO4•− signals became weaker due to the decreased amount of PS (Figure 5b). Moreover, an approximately 1:2:2:1 quartet pattern (aH= aN = 14.9 G) was observed, which was assigned to the DMPO−•OH adduct,43 confirming that surface OH group on Al2O3 promotes the conversion of SO4•− into •OH (eq 10). After extracting the milled sample with moisture, formation of the DMPO-•OH adduct was observed in the two mixtures and MC−Al2O3−PS exhibited a stronger DMPO−•OH signal than MC−PS (SI Figure S12), thereby supporting the conversion of SO4•− to •OH on the Al2O3 surface (•OHsurf). Moreover, the ESR signals for the milled samples of PS−DMPO and Al2O3− PS−DMPO remain considerably intense after the milled samples were placed at room temperature for 22 and 4 h, respectively (SI Figure S13). The long-lived DMPO-radical adducts on the milled samples are similar to “environmentally persistent free radicals” that formed on the surface of engineered nanomaterials during combustion or thermal treatments,44 probably because the association of the free radical formed in situ with the surface of milled solids stabilizes the radical.

S2O82 −surf ⎯⎯⎯⎯⎯⎯→ 2SO4•−surf

(9)

OH−surf + SO4•−surf →• OHsurf + SO4 2 −

(10)

•−

To evaluate the contributions of SO4 surf and •OHsurf, the effect of surface OH on Al2O3 was studied. We preheated Al2O3 powders at 100 °C for 24 h. This preheating decreased the surface water and/or OH− on the Al2O3 by approximately 50% and consequently the defluorination efficiency of PFOA in MC−Al2O3−PS after 2 h milling decreased from 100% to 64% (SI Figure S14). The MC degradation of PFOA was further performed by using 18O-labeled Al2O3 as a comilling agent. Prior to the degradation experiments, Al2O3 powders were dispersed in H218O for 1 h to ensure full oxygen exchange between H218O and the surface OH on Al2O3, following by vacuum drying at 60 °C for 1 h to remove the surface water on Al2O3. For convenience, the obtained 18O-labeled Al2O3 was named 18OH−Al2O3. It was found that the defluorination efficiency of PFOA was reduced from 100% in MC−Al2O3−PS to 70.6% in MC−(18OH−Al2O3)−PS at a 2 h milling (Figure 6a). Meanwhile, the isotope distribution of 18O in another dominant product, CO2, was significantly changed. As shown in Figure 6b, the collected gaseous product shows characteristic FT−IR absorptions of νas(C16O2) at 2361 and 2337 cm−1 in MC−PFOA−Al2O3−PS. These two peaks not only shift to lower wavenumbers by approximately 17 cm−1 but also weaken in MC−PFOA−(18OH−Al2O3)−PS at the same milling time, thus indicating that 16O in CO2 exchanged by 18O and the production of CO2 is slower than in MC−PFOA−Al2O3−PS. Accordingly, the mass spectrum of CO2 produced in MC− PFOA−Al2O3−PS exhibits a dominant m/z of 44 (C16O16O, 99.5%), being similar to natural isotope abundance of CO2 in 8309

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Environmental Science & Technology Scheme 1. Schematic Representation for MC Degradation of PFOA in the Presence of Both Al2O3 and PS

to-solid reactions by producing a transient temperature and initiating the rotation, alignment and reorientation of molecules. On one hand, ball milling induces the decarboxylation reaction of PFOA, thereby yielding CO2 and an unstable perfluoroheptyl anion (−C7F15) on the Al2O3 surface (eq 11) which preferentially reacts with the proton of an OH group on Al2O3 to yield C7F15H (eq 12). On the other hand, the high-energy impact facilitates the decomposition of PS into SO4•−, which consequently oxidizes OH−surf on Al2O3 to •OHsurf. As demonstrated in Figures 5 and 6, the active species for the degradation of PFOA is •OHsurf in MC−Al2O3−PS. Hydrogen abstraction from C7F15H by •OHsurf is followed by the formation of •C7F15 (eq 13), which couples with •OHsurf to yield C7F15OH (eq 14). C7F15OH transforms into C6F13COF through the elimination of HF (eq 15) and reacts with the surface water to yield C6F13COOH together with H+ and F− (eq 16). Similarly, C6F13COOH is bound on the Al2O3 surface via the COO−···Al and Al···F interactions, which activates the C−COO bond and C−F bonds, respectively (Scheme 1). This leads the decarboxylation, which yields − C6F13, to take place as a second reaction on the Al2O3 surface.

air (99.6%) and obtained from the reaction of Na2CO3 with HNO3 (99.3%) (Figures 6c and SI Figure S15). In contrast, the ratios of CO2 with m/z of 44 (C16O16O), 46 (C16O18O), and 48 (C18O18O) in MC−PFOA−(18OH−Al2O3)−PS were changed to 63.6%, 32.0%, and 4.4%, respectively (Figure 6c). Because the background influence of air may cause a negative deviation in the evaluation of the absolute isotope ratio of 18 O/16O (R18O/16O), we did not calculate R18O/16O. However, the mass area ratio of C16O18O to C18O18O was 7.3, which suggests that one of the two oxygen atoms in CO2 comes from 18 O-labeled surface OH on Al2O3. After collecting CO2 gas in NaOH solution and converting it to MnCO3 precipitant, the stable isotope ratio of 18O/16O (δ18O, vs Vienna standard mean ocean water (VSMOW) was measured on an elemental analyzer coupled to an isotope ratio mass spectrometer. The δ18O values for the MnCO3 obtained from MC−PFOA− Al2O3−PS and the two reference compounds CO and refMnCO3 were 16.08‰, 9.07‰, and 38.02‰, respectively, all of which are close to those of Pee Dee Belemnite, a carbonate isotope standard (30‰ vs VSMOW, SI Table S4). In contrast, the δ18O of MnCO3 obtained from MC−PFOA−(18OH− Al2O3)−PS was 3251.50‰, which is 154 times higher than the average value of the above-mentioned three controls (21.05‰). The R18O/16O values of CO32− and its precursor CO2 in the case of MC−PFOA−(18OH−Al2O3)−PS were calculated to be 0.32 and 0.61, respectively (SI Table S4). These results clearly demonstrated that the O atom in the surface OH on Al2O3 can be incorporated into PFOA and finally transformed into CO2. By combining the fact that the conversion of PFOA to CO2 in MC−Al2O3 and MC−PS was less than 20%, it is reasonable to propose that the surface OH− on Al2O3 is oxidized by SO4•−surf to •OHsurf, which is then predominant species in the mineralization of PFOA. Reaction Mechanism. Based on the above results, the reaction mechanism of degradation of PFOA using both Al2O3 and PS under the conditions of ball milling is proposed in Scheme 1. After mixing PFOA with one agent, PFOA was tightly bound to Al2O3 through the COO−···Al bond and > AlOH···F−C hydrogen bonding (eqs 4 and 5), whereas the O2− of PS acted as the proton acceptor to bind PFOA through the hydrogen bond in MC−PS. In the ternary mixture, the interaction between PFOA and Al2O3 does not change, but Al2O3 replaces PFOA to bind PS via hydrogen bonding of > AlOH···O2− (PS) (eq 6). That is, Al2O3 strongly anchors the PFOA and PS (Scheme 1). During the MC treatment, the high-energy impact has multiple roles in the activation of solid-

milling

C7F15COO− ··· Al < ⎯⎯⎯⎯⎯⎯→ −C7F15 ··· Al < + CO2

(11)

(−C7F15)surf + H+surf → (C7F15H)surf

(12)

(C7F15H)surf + •OH surf → (•C7F15)surf + H 2O

(13)

•

•

( C7F15)surf + OHsurf → (C7F15OH)surf

(14)

(C7F15OH)surf → (C6F13COF)surf + (HF(H+ + F−))surf (15)

C6F13COFsurf + H 2Osurf → C6F13COOHsurf + (HF(H+ + F−))surf

(16)

The successive degradation including decarboxylation, protonation, hydrogen-abstraction, radical coupling, elimination of HF, and hydrolysis (eqs 11−16), occur to yield shortchain PFCA products until CF3COOH is reached, which degrades to −CF3 on the Al2O3 surface (eq 11). Degradation and defluorination are repeated to yield CO2, H+, and F−as the final products (eqs 12−15). Thus, complete degradation and defluorination are achieved. It is noted that the mineralization of 1 mol PFOA consumes 14 mol •OHsurf, of which 7 mol •OHsurf are converted to CO2, and the others to H2O (SI Scheme S1). Moreover, the mineralization of 1 mol PFOA 8310

DOI: 10.1021/acs.est.9b00486 Environ. Sci. Technol. 2019, 53, 8302−8313

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Environmental Science & Technology Table 2. Comparison of Energy Consumption for the Degradation of PFOA by Different Treatment Methods treatment method MC−Al2O3−PS MC−KOH MC−NaOH UV/In2O3 UV/BiOCl UV/PS UV/H2O2 US+SnO2−Sb/carbon aerogel electrode

conditions

power (W)

deflurinationefficiency (%)

energy (kJ μmol−1)a

ref

0.604 mmol PFOA,2.58 g Al2O3, 1.37 g PS 0.483 mmol PFOA,4.6 g KOH 0.483 mmol PFOA,4.6 g NaOH 0.1 mM PFOA, 0.4 L0.5 g L−1 In2O3 0.02 mM PFOA, 0.2 L,0.5 g L−1 BiOCl 29.6 μmol PFOA,1.10 mmol PS 0.02 mM PFOA, 0.2 L,40 mM H2O2, 4 mM Fe2+ 0.1 mg L−1 PFOA, 0.06 L,immersed area 5 cm2

750 750 750 23 10 200 9 >50c

100 (2 h) 97 (4 h) 19 (4 h) 33.7 (4 h) 59.3 (24 h) 59.1 (4 h) 46 (24 h) 81 (5 h)

0.15b 0.39b 1.99b 1.64 24.3 10.8 28.2 >5.10

this work 21 21 49 50 45 51 52

Energy consumed to release one micromolar F− ions from the degradation PFOA, which is calculated from SI eq S11 (SI Text S3). bPlanetary ball mill is a robust floor model with four grinding stations, and thus the treated PFOA should be 4 times of the added pollutants. cThe total power includes the ultrasonic irradiation (50 W) and the employed power in electrochemical experiment, but the latter of them is unknown. a

nation.25 In MC−PS, SO4•− oxidizes the PFOA adsorbed on the surface of PS powders to form unstable •C7F15 through a one-electron oxidation, deprotonation, and decarboxylation (SI Text S2, eqs S1−S3). Further reactions between •C7F15 and PFOA yield C7F15COOC7F15, C7F15COC7F15, C7F15OC7F15, and C7F15COC6F13 as the main products (SI Text S2, eqs S4− S7, Figure S7). Thus, the TOC removal and defluorination ratios were as low as 11% and 6.4%, respectively (Figures 1b and 2b), although 61% of the PFOA was degraded in MC−PS after a 2 h milling. It is also noted that the defluorination efficiency of PFOA in MC−PS (70%) induced by photolysis or thermolysis of PS in aqueous solution,45−47 due to the different degradation pathways of PFOA. Upon treatment with PS in water, SO4•− can be converted into •OH, and both of them contribute to the degradation of PFOA. The unstable •C7F15 generated from oxidation by SO4•− usually reacts with •OH, dissolved O2 and/or sufficient H2O to yield C6F13COOH, which follows an analogous mechanism to produce short-chain PFCAs products until complete mineralization and defluorination occurs.45−48 The fast diffusion of chemical species in water will not cause PFCA molecules to remain close to the PS species until the parent PFCA molecule is completely defluorinated. As a result, quite large amounts of short-chain PFCAs were detected as major products during the defluorination of PFOA in PSinvolved aqueous solution.45−48 Applications and Implications. We found complete degradation and defluorination of PFOA through activation of C−COO− and C−F bonds and generation of •OH by the MC treatment using Al2O3 and PS as comilling agents, respectively. The MC treatment was also used to degrade the homologues of PFOA such as PFPA, PFHxA, and PFHpA under the same conditions (i.e., mAl2O3/mPS/mPFCA, 2.58:1.37:0.25; mb/ms, 50:1). After ball milling for 2 h, such PFOA homologues are degraded with a defluorination efficiency of 100% (SI Figure S17). Perfluorooctanesulfonic acid (PFOS) is another important perfluorinated pollutant. In general, PFOS is much more resistant to oxidative degradation by SO4•− than PFCAs, and hence, no transformation of PFOS was observed in aqueous solution or soil slurry in the presence of heat-activated PS.47,48 We used the developed MC method to degrade PFOS, and found that 56% of the added PFOS was slowly degraded in MC−Al2O3−PS with a defluorination efficiency of 9.7% after a 2 h milling (SI Figure S18). The considerably slower degradation of PFOS than PFOA was in good agreement with the above-mentioned opinion that PFOS is much more

generates 8 mol CO2, of which 7 mol come from the produced short-chain PFCAs (SI Scheme S1). Each short-chain PFCA molecule (CnF2n+1COOH) contains two external oxygen atoms: one from the addition of •OHsurf (eqs 14 and 15, SI Scheme S1), and the other from surface water via the hydrolysis reaction (eq 16, SI Scheme S1). In 18O-labeling experiments, all the surface OH− groups on Al2O3 were exchanged by the 18OH of H218O, followed by drying at 60 °C for 1 h to remove surface water on the Al2O3. During the MC treatment, the oxidation of surface 18OH− on Al2O3 by SO4•−surf yields •18OHsurf, while surface water comes from moisture at natural isotope abundance (H216O), giving a theoretical R18O/16O in the produced CO2 of 0.78 as shown in eq 17. The expected observations of C16O18O as a main product and an R18O/16O(CO2) = 0.6 (Figures 6b and 6c) clearly demonstrate that the MC degradation of PFOA in MC− Al2O3−PS follows a •OHsurf involved pathway (eqs 11−16). C7F15C00H 0 + 718OH−surf + 7H+surf + 14·18OHsurf + 7H 216Osurf → CO2 + 7C16O18O + 7H 218O + 7H 2 + 15HFsurf

(17)

17 From eq 17, a complete defluorination of 1 mol PFOA may yield 15 mol F− and consumes 14 mol •OHsurf, being equivalent to 7 mol PS as calculated from eqs 9 and 10. According to eq 3, the utilization efficiency of PS (ηPS) was calculated from the ratio of the stoichiometrically required PS for generating F− to the actual consumption. For the MC degradation of PFOA (0.25 g) at nAl2O3/nPS/nPFOA of 42:8.4:1, the residual PS decreased to 3.7% when the defluorination efficiency of PFOA was 100% (SI Figure S16a). The value of η PS in MC−Al 2 O 3 −PS remained at 87% during the defluorination of PFOA (SI Figure S16b). This value is much higher than those for the defluorination of PFOA in aqueous solution (8∼22%) and aqueous soil slurry (