Complete defluorination and mineralization of perfluorooctanoic acid

Subscriber access provided by Bethel University

Remediation and Control Technologies

Complete defluorination and mineralization of perfluorooctanoic acid by mechanochemical method using alumina and persulfate Nan Wang, Hanqing Lv, Yuqi Zhou, Lihua Zhu, Yue Hu, Tetsuro Majima, and Heqing Tang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.9b00486 • Publication Date (Web): 31 May 2019 Downloaded from http://pubs.acs.org on June 4, 2019

Just Accepted “Just Accepted” manuscripts have been peer-reviewed and accepted for publication. They are posted online prior to technical editing, formatting for publication and author proofing. The American Chemical Society provides “Just Accepted” as a service to the research community to expedite the dissemination of scientific material as soon as possible after acceptance. “Just Accepted” manuscripts appear in full in PDF format accompanied by an HTML abstract. “Just Accepted” manuscripts have been fully peer reviewed, but should not be considered the official version of record. They are citable by the Digital Object Identifier (DOI®). “Just Accepted” is an optional service offered to authors. Therefore, the “Just Accepted” Web site may not include all articles that will be published in the journal. After a manuscript is technically edited and formatted, it will be removed from the “Just Accepted” Web site and published as an ASAP article. Note that technical editing may introduce minor changes to the manuscript text and/or graphics which could affect content, and all legal disclaimers and ethical guidelines that apply to the journal pertain. ACS cannot be held responsible for errors or consequences arising from the use of information contained in these “Just Accepted” manuscripts.

is published by the American Chemical Society. 1155 Sixteenth Street N.W., Washington, DC 20036 Published by American Chemical Society. Copyright © American Chemical Society. However, no copyright claim is made to original U.S. Government works, or works produced by employees of any Commonwealth realm Crown government in the course of their duties.

Page 1 of 37

Environmental Science & Technology

1

Complete defluorinationand mineralization of perfluorooctanoic acid by a

2

mechanochemical method using alumina and persulfate

3

Nan Wanga,+, Hanqing Lva,+, Yuqi Zhou a, Lihua Zhua,*,Yue Hub, Tetsuro Majima a, Heqing

4

Tangb,*

5

a

6

Technology, Wuhan 430074, P. R. China

7

bCollege

8

430074, P. R. China

9

+These

College of Chemistry and Chemical Engineering, Huazhong University of Science & of Resourcesand Environmental, South-Central University for Nationalities, Wuhan

authors contributed equally to this work.

10

ACS Paragon Plus Environment

1


12

ABSTRACT: Perfluorooctanoic acid (PFOA) is a persistent organic pollutant that has

13

received concerns worldwide due to its extreme resistance to conventional degradation. A

14

mechanochemical (MC) method was developed for complete degradation of PFOA by using

15

alumina (Al2O3) and potassium persulfate (PS) as comilling agents. After ball milling for 2 h,

16

the MC treatment using Al2O3 or PS caused conversion of PFOA to either

17

1-H-1-perfluoroheptene or dimers with a defluorination efficiency lower than 20%, but that

18

using both Al2O3 and PS caused degradation of PFOA with a defluorination of 100% and a

19

mineralization of 98%. This method also caused complete defluorination of other C3~C6

20

homologues of PFOA. The complete defluorination of PFOA attributes to Al2O3 and PS led to

21

the weakening of the C-F bond in PFOA and the generation of hydroxyl radical (•OH),

22

respectively. During the MC degradation, Al2O3 strongly anchors PFOA through COO−-Al

23

coordination and in-situ formed from Lewis-base interaction and PS through hydrogen bond.

24

Meanwhile, mechanical effects induce the homolytic cleavage of PS to produce SO4•−, which

25

reacts with OH group of Al2O3 to generate •OH. The degradation of PFOA is initiated by

26

decarboxylation as a result of weakened C-COO− due to Al3+ coordination. The subsequent

27

addition of •OH, elimination of HF, and reaction with water induce the stepwise removal of

28

all carboxyl groups and F atoms as CO2 and F−, respectively. Thus, complete defluorination

29

and mineralization are achieved.

30


2

Page 2 of 37

Page 3 of 37

32


TOC Art

33


3


35

Page 4 of 37

INTRODUCTION

36

Perfluorooctanoic acid (C7F15COOH, PFOA), perfluorooctane sulfonate (C7F15SO3H,

37

PFOS), and other fluorosurfactants have been widely applied in textiles, paper and packaging

38

materials, cookware, and fire-fighting foam productions. Due to their disposal, these

39

fluorocarbons are frequently detected in aquatic environments,1 soils, sludge and sediments.2

40

Because PFOA and PFOS are persistent organic pollutants (POPs),3 their elimination is

41

necessary.

42

PFOA is thermally and chemically stable because of high dissociation energy of C−F

43

bonds (533 kJmol−1).4 To induce the degradation of PFOA in aqueous media, several

44

chemical approaches have been explored, such as UV-sulfite reduction,5,

45

electro-Fenton oxidation,7,8 photocatalytic oxidation,9 and sulfate radical anion (SO4•−)-based

46

oxidation.10-12 However, few methods have been developed for the degradation of PFOA in

47

solids. Incineration at temperatures above 500 °C is a practical method to treat organic

48

pollutants in solid, although the thermal treatment of PFOA produces undesirable greenhouse

49

gases and corrosive gases such as CF4, C2F4, and HF.13 Thus, development of complete and

50

environment-friendly disposal technologies is urgently needed for treating PFOA in solids.

6

photo- and

51

Mechanochemical (MC) degradation is a promising method for the disposal of organic

52

pollutants in solid. In an MC treatment, the degradation reaction usually begins with the MC

53

activation of milling agents by particle crushing, build-up of structural defects, rupture of

54

bonds, heating, etc., under intensive mechanical stresses. The efficiency of the MC

55

degradation of organic pollutants is improved by using appropriate additives. For example,

56

CaO14, CaO-SiO215, Fe16, zero-valent metal (e.g., Fe, Al)-SiO217, persulfate (PS)18, and

57

PS-CaO19 have been used as efficient additives in MC degradation of chlorinated and

58

brominated organic pollutants, although these additives are not efficient enough individually

59

for the MC degradation of PFOA (Table S1). Yan et al. investigated the MC degradation of a


4

Page 5 of 37


60

chlorinated polyfluorinated ether sulfonate (F-53B, a PFOS alternative) by using sodium

61

persulfate and NaOH as comilling agents, and found that the defluorination efficiency of

62

F53-B was 54% after ball milling for 8 h.20 Zhang et al. reported that ball milling for 8 h

63

produced defluorination efficiency of PFOA and F-53B greater than 80% by adding KOH as a

64

milling agent with a KOH/PFOA molar ratio of 170:1, but less than 20% by adding NaOH.21,

65

22

66

The relatively poor defluorination in the MC degradation of PFOA is closely related to

67

the chemical inertness of C−F bonds. Activation of a C−F bond and the following cleavage

68

involve the formation of thermodynamically favorable bonds such as H−F, B−F, Al−F, Si−F,

69

and Ge−F bonds.23, 24 Recently, we have found that the mechanocaloric effect promotes the

70

dehydration of the Al2O3 surface to produce coordinatively unsaturated Lewis acid sites,

71

inducing the activation of C−F bonds to convert PFOA to a valuable product,

72

1-H-1-perfluoroheptene.25 Cagnetta et al. reported that the use of La2O3 as a milling agent

73

converted perfluorinated compounds into LaOF, a luminescent material.26 This MC method is

74

preferable for the degradation of solid PFOA, because not only is PFOA degraded, but

75

organo-fluorine blocks in PFOA are also recovered for synthesizing polyfluoroalkenes or

76

LaOF. However, MC treatment is not appropriate for degradation of PFOA at ratios lower

77

than 20% in solids because the economy of this method becomes poor for recovery of the

78

fluorine resource. In most cases, the PFOA content is too low to be recovered as a resource.

79

Therefore, a degradation of PFOA with complete defluorination in solids to harmless

80

substances is desirable.

81

We recently have reported using the MC method with PS to cause the homolytic

82

cleavage of the peroxide bond in PS to produce SO4•−, inducing the decomposition of

83

brominated organic pollutants.18 Considering the activation effect of Al2O3 on C−F bonds of

84

PFOA and generation of SO4•− from the dissociation of PS by the MC treatment, we


5


85

anticipated that Al2O3 and PS might lead to complete MC degradation and defluorination of

86

PFOA. In the present work, we developed an MC method for degradation of PFOA by using

87

both Al2O3 and PS as comilling agents, and clarified the reaction mechanism.

88



89

Chemicals and Reagents. Perfluorooctanoic acid (PFOA, C7F15COOH, 98% in purity) was

90

purchased from J&K China Chemical, Beijing, China. Shorter chain perfluorocarboxylic

91

acids (PFCAs) including perfluoroheptanoic acid (PFHpA, >98%), perfluorohexanoic acid

92

(PFHxA, >98%) and perfluoropentanoic acid (PFPA,>98%) were purchased from Tokyo

93

Chemical Industry, Shanghai, China. Alumina (Al2O3, 99.9%), 5,5-dimethyl-1-pyrroline

94

(DMPO, analytical grade), and diphenyl-1-picrylhydrazyl (DPPH•, 97%) was obtained from

95

Aladdin, Shanghai, China. H218O with an isotope abundance of 98% was purchased from

96

Jiangsu Huayi Isotopes Chemical Limited. Potassium persulfate (PS, analytical grade) was

97

obtained from Sinopharm Chemical Reagent Co., Ltd., Shanghai, China. All solvents used for

98

extraction and high-performance liquid chromatography (HPLC) analysis were analytical

99

grade and HPLC grade, respectively. Milli-Q water with conductivity of 18.2 MΩ cm (Merck

100

EXPERIMENTAL SECTION

Millipore GmbH, Germany) was used in the present work.

101

Degradation Experiments. The MC degradation of PFOA was conducted in a planetary

102

ball mill at room temperature (BM4, Beijing Grinder Instrument, China) by using Al2O3

103

and/or PS as milling agents. Typically, 0.25 g PFOA and 3.95 g milling agents (PS, Al2O3,

104

and/or their mixtures) were mixed in the stainless steel pot (250 mL) of the ball milling

105

machine, which was then filled with 18 stainless steel balls with a diameter (d) of 6 mm and

106

10 balls with d = 10 mm (the total mass of balls was 210 g). When PS and Al2O3 were used as

107

comilling agents, the molar ratio of Al2O3/PS ( nAl O /nPS ) was set at 5 unless otherwise stated.

108

The mass of the total milled sample was 4.2 g, and the ball-to-sample mass ratio was fixed at

109

50:1. After the milling materials were well predistributed in a pot, the pot was sealed tightly

2

3


6

Page 6 of 37

Page 7 of 37


110

and the planetary ball mill was operated at 350 rpm under atmospheric conditions, with

111

automatic change in the direction of rotation every 15 min. At given time intervals, the ball

112

milling was stopped and solid samples were taken out for chemical analysis, unless specified

113

elsewhere. Replicate runs (n ≥ 3) were carried out for each test, and the relative

114

standard deviations were less than 5%.

115

For isotopic labeling experiments, both the PS and Al2O3 were preheated at 100 °C for

116

24 h and kept in a desiccator prior to MC degradation experiments. H218O (1 mL) was mixed

117

with Al2O3 (10 g) in a desiccator for 1 h to ensure full oxygen exchange between H218O and

118

surface OH on the Al2O3, followed by vacuum drying at 60 °C for 1 h. Then,

119

Al2O3 powders were used for the MC treatment of PFOA.

120

Analysis. Each milled sample (0.020g) was extracted with ethanol (3 times, for a total of 10

121

mL) under ultrasound irradiation (KQ-200KDE, Kunshan Ultrasonic Instruments, China) for

122

10min. After removing the solids by centrifugation, the collected solution was filtered through

123

a 0.22 μm membrane and then subjected to HPLC analysis. Similar to our previous work,27

124

the PFOA concentration was measured with HPLC on an Ultimate 3000 series system

125

(Dionex, Idstein, Germany) equipped with a Corona Ultra RS charged aerosol detector (CAD,

126

Thermo Scientific, Bellefonte, PA, USA). The analysis was performed on a Symmetry C8

127

column (150 mm × 3.9 mm; 5 μm particle size; Waters, Ireland). Column oven temperature

128

was set at 30 °C. The mobile phase was 30% ammonium acetate (5 mmol L−1 with pH

129

adjusted to 5.0 ± 0.2 by acetic acid) and 70% methanol by volume, the flow rate was 1 mL

130

min−1, and the injection volume was 40 μL. The degradation efficiency of PFOA (DGEPFOA)

131

is calculated from eq. 1,

132

133

  n DGE PFOA  1  PFOA,t   100%  nPFOA,0 

18O-labeled

(1)

where nPFOA,0 and nPFOA,t are moles of PFOA at reaction times 0 and t, respectively. In


7


Page 8 of 37

134

addition, the ethanol-extracted solutions were analyzed using an 1100 LC/MSD trap

135

system (Agilent, USA) with an electrospray ionization (ESI) source to identify possible polar

136

intermediates.

137

To identify volatilized organic intermediates, both products in the reactor and gaseous

138

products in the head space of the reactor were collected by using acetone as a solvent. The

139

extract was filtered through a 0.22 μm membrane and then directly analyzed by GC/MS

140

(Thermo Fisher, USA) under the same conditions reported in our previous work.25 For the

141

analysis of the CO2 produced, the gases were collected in 20 mL NaOH (0.5 mol L−1) after the

142

ball milling. The CO2 absorbed in NaOH was converted into CO32− and then detected by ion

143

chromatography (IC) on a Dionex ICS-1500 equipped with a CD 25 conductivity detector,

144

IonPac AS 23 column and KOH as the eluent (28 mmol L−1, 1.0 mL min−1). In the isotopic

145

labeling experiments, the produced CO2 was analyzed using Fourier transform infrared

146

(FT-IR) spectrometry (Bruker, Germany) by introducing the gaseous products into the gas

147

cell. The distribution of 18O in CO2 was analyzed not only by GC/MS (Agilent 7890B-5977A

148

GC/MSD) with a DB-5MS column (30 m × 0.25 mm; 0.25 μm particle size) and quadruple

149

mass analyzer, but also by isotope ratio mass spectrometry after converting the gaseous CO2

150

to a precipitate of MnCO3. After the MC reaction, the generated CO2 was absorbed by a

151

NaOH solution, followed by isolating the milled powders using centrifugation and filtration.

152

The filtrate, with a pH approximately 7.5, was adjusted to pH 11 with NaOH (0.5 mol L−1),

153

followed by adding MnSO4•H2O (0.08 g) to yield a MnCO3 precipitate. The

154

abundance in MnCO3 was analyzed by using an elemental analyzer (EA IsoLink CN/OH,

155

Thermo Fisher USA) coupled to an isotope ratio mass spectrometer (Delta V Advantage,

156

Thermo Fisher). In the reaction furnace of the elemental analyzer at 1380 °C, the evolved

157

oxygen in MnCO3 reacted with carbon to yield CO, which was separated and then transported

158

into an isotope ratio mass spectrometer. Because the conversion of MnCO3 to CO strongly


8

18O

isotope

Page 9 of 37


159

depends on temperature, two reference compounds including CO and MnCO3 that obtained

160

from the reaction of Na2CO3 with MnSO4 (named ref-MnCO3) were subjected to the same

161

measurement conditions.

162

To determine the amount of F− released from PFOA, the milled samples (20 mg) were

163

dispersed in 10 mL water and 0.5 mL aqueous ammonia. The suspensions were centrifuged

164

and filtered through a 0.22 μm membrane. The filtrate then was diluted to 50 mL with 10 mL

165

total ionic strength adjustment buffer (TISAB, consisting of HAc-NaAc and 10 g L−1 sodium

166

citrate) and water. Finally, F− was measured with a F− selective electrode. The recovery yields

167

of F− were greater than 95% as determined using the standard addition method by spiking NaF

168

(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)

169

(Table S2). The defluorination efficiency of PFOA (DFEPFOA) is evaluated from eq. 2,

170

DFE PFOA 

ndetected F ,t nF,0

 100% 

ndetected F ,t nPFOA,0  15

 100%

(2)

171

where ndetected F ,t and nF,0 are moles of F− detected experimentally and F calculated

172

from the initial PFOA (mol), respectively.

173

The

residual

PS

in

the

milled

samples

was

determined

with

a

KI

174

spectrophotometric method.28 Prior to the measurement, the milled samples (20 mg) were

175

dispersed in 10 mL water under ultrasound (KQ-200KDE, Kunshan Ultrasonic Instruments,

176

China) for 10 min. After centrifugation and filtration through a 0.22 μm membrane, 1.5 mL

177

filtrate was mixed with 1.0 mL phosphate buffer (pH 6.86) and 0.5 mL KI (2 M). After 15

178

min, 0.20 mL of this mixture was diluted to 10 mL with water and then measured at 355 nm

179

on a UV-vis spectrophotometer (Cary 50, Varian, USA). The recovery yields of PS were

180

greater than 98% with the standard addition method by spiking PS (1.37 g) into a mixture of

181

PFOA (0.25 g) and Al2O3 (2.58 g). The utilization efficiency of PS (PS) is defined as the ratio

182

of the stoichiometric requirement to the actual consumption, which is calculated from the


9


Page 10 of 37

183

moles of generated F− and consumed PS ( ndetected F ,t and nPS, 0  ndetected PS, t , respectively) (eq.

184

3),

 PS 

185

ndetected F



,t

nPS, 0  ndetected PS,t

 100%

(3)

186

where ndetected PS, t

and nPS,0 are moles of PS detected experimentally and calculated

187

from the initial PS, respectively, and  is the stoichiometric ratio of the moles of

188

generated F− to the moles of consumed PS.

189

When necessary, solid samples were taken at specified time intervals during the ball

190

milling and then were characterized using spectroscopic techniques. FT-IR spectra were

191

measured on an Equinox 55 (Bruker, Germany) with the KBr disk method from 400 cm−1 to

192

4000 cm−1. X-ray photoelectron spectrometry (XPS) was conducted on Kratos AXIS-ULTRA

193

DLD-600W (Shimadzu, Japan), in which the binding energy is calibrated using the C 1s peak

194

as 284.6 eV. 19F and 27Al MAS NMR experiments were performed at room temperature on a

195

Bruker Avance III 500 MHz spectrometer (Bruker, Germany) equipped with a 2.5 mm Bruker

196

1H/19F/X

197

kHz.

probe. The samples were loaded into a zirconia rotor and spun at an MAS rate of 20

198

A Bruker EMX-nano electron spin resonance (ESR) instrument was employed to detect

199

radicals with DMPO as the spin trapping reagent. Prior to measurements, a small amount of

200

DMPO instead of PFOA was added to the milling mixture. After being milled for 15 min at

201

300 rpm, a portion of the milled sample (ca. 0.1 g) was directly collected and another portion

202

(20 mg) of the milled mixture was dispersed in 10 mL water. Both the solid and

203

water-extracted samples were measured under the following conditions: center field, 3480 G;

204

sweep width, 200 G; microwave frequency, 9.77 GHz; modulation frequency, 40 kHz; power,

205

10 mW; scan times, 3.

206

 RESULTS ANDDISCUSSION


10

Page 11 of 37


207

Complete Degradation and Defluorination of PFOA using Mechanochemistry. Figures 1a

208

and 1b illustrate the MC degradation and defluorination of PFOA (0.1 g) in kaolin soil (1.5 g)

209

with PS, Al2O3 or a mixture of PS and Al2O3 as milling agents (5.7 g). After milling for 2 h,

210

all the added PFOA in the three cases was nearly degraded, but defluorination efficiencies

211

calculated from the generated F− were 10%, 27% and 85% in MC-Al2O3, MC-PS and

212

MC-Al2O3-PS, respectively. When Al2O3 (2.4 g) and PS (3.3 g) were divided into three equal

213

portions and added for the three reaction times of 0, 30, and 60 min, the degradation and

214

defluorination extents of PFOA achieved 99% and 96% at 2 h, respectively. This indicates

215

that the use of both Al2O3 and PS favors the MC defluorination of PFOA. Because the soil

216

matrix will bring more difficulties in clarifying the roles of Al2O3 and PS, the subsequent

217

degradation experiments were carried out using pure PFOA as a model pollutant.

Degradation efficiency /%

75

50

PS Al2O3+PS

25

0

218

Al2O3

Al2O3+PS(1/33)

0

40 80 Milling time /min

Defluorination efficiency /%

100 (b)

100 (a)

75 Al2O3 PS Al2O3+PS

50

Al2O3+PS(1/33) 25

0

120

0

30


120

219

Figure 1. Time profiles of degradation (a) and defluorination (b) of PFOA (0.1 g) in kaolin

220

soil (1.5 g) under 350 rpm and a ball-to-sample mass ratio (mb/ms) of 50 in the presence of

221

Al2O3 (5.7 g), PS (5.7 g), or the mixture of Al2O3 (2.4 g) and PS (3.3 g). All milling agents

222

were added in one time before milling, except for the system of Al2O3+PS (1/3  3), in which

223

PS and Al2O3 were divided into three equal portions and added at reaction time of 0, 30, and

224

60 min.

225

Figures 2a and 2b compare the MC degradation and defluorination of powdered PFOA


11


Page 12 of 37

226

(0.25 g) by using 3.95 g milling agents. After a 2 h milling with either PS ( nPS /nPFOA = 24) or

227

Al2O3 ( nAl O /nPFOA = 64), the degradation efficiency of PFOA was greater than 60%, but the 2

3

228

defluorination extent was lower than 20%. When using comilling agents at nAl O /nPS = 5,

229

both degradation and defluorination efficiencies reached 100% after 2 h milling. Moreover, at

230

any time, the defluorination extent of PFOA was close to the degradation in MC-Al2O3-PS,

231

thus confirming that successive degradation and defluorination continued until there is a

232

complete removal of F atoms. In addition, the fate of fluorine was identified by solid-state

233

nuclear magnetic resonance (SS NMR) and XPS measurements. High-resolution XPS spectra

234

of F 1s and NMR spectra of

235

MC-Al2O3-PS is bonded to the Al2O3 surface to form Al−F bond (Text S1, Figures S1 and

236

S2). This provides additional merit for reducing the secondary pollution because the bonding

237

of F− on Al2O3 reduces the dissociation of F− into water and thus avoids the possible

238

promotion effect of bone disease from excess F−.

2

19F

27Al

and

3

demonstrated that F− eliminated from PFOA in

239

Figure 2c shows the effect of nAl O /nPS on the degradation and defluorination ratios of

240

PFOA after a 2 h MC treatment. As nAl O /nPS increases from 0 to 2, the degradation extent

241

of PFOA increases from 61% to 96%. A further increase in nAl O /nPS to 5 results in a

242

degradation ratio for PFOA of 100%. However, a different volcano plot is observed for the

243

effect of nAl O /nPS on the defluorination efficiency of PFOA. The defluorination ratio of

244

PFOA increases from 6.2% to 98% when nAl O /nPS increases from 0 to 3 and reaches 100%

245

at nAl O /nPS = 5. After that, the defluorination ratio of PFOA significantly decreases to 42% as

246

nAl O /nPS increases to 10. The use of Al2O3 alone leads to a defluorination ratio of 18%, even

247

though all the added PFOA is removed. The above observations are related to different roles

248

of comilling agents on the degradation and defluorination of PFOA. Al2O3 exhibits better

2

3

2

3

2

2

3

2

2

2

3

3

3

3


12

Page 13 of 37


249

performance than PS on the degradation of PFOA via a decarboxylation process as a result of

250

weakened C-COO− due to Al3+ coordination (Figure 2a), leading to that a larger nAl O /nPS

251

caused a higher degradation ratio of PFOA. However, neither the in situ generated electrons

252

in MC-Al2O3 nor the generated SO4•− in MC-PS was efficient in the defluorination of PFOA

253

(Figure 2b). The complete defluorination of PFOA is attributed to Al2O3 and PS, which

254

caused the weakening of the C-F bond in PFOA and the generation of hydroxyl radicals

255

(•OH), respectively. During ball milling, mechanical effects induce the decomposition of PS

256

into SO4•−, which reacts with surface OH− (OH−surf) on Al2O3 to generate •OHsurf. As

257

nAl O /nPS decreases, the total amount of OH−surf on Al2O3 decreases. In contrast, as nAl O /nPS

258

increases, the amount of PS decreases. Either of them would result in the insufficient

259

generation of •OHsurf from the reaction of SO4•− with OH−surf. Thus, a complete defluorination

260

of PFOA requires a moderate value of nAl O /nPS .

2

2

2

3

2

3

3

3

261

Figure 2d illustrates the effect of the ball-to-sample mass ratio (mb/ms) on the MC

262

degradation of PFOA at a fixed nAl O /nPS /nPFOA of 42: 8.4: 1. As mb/ms increases from 15 to

263

50, both the degradation and defluorination ratios of PFOA at 2 h increases from

264

approximaterly 35% to 100%. However, slight increases in the degradation and defluorination

265

ratios of PFOA are observed with an increase in mb/ms to 75 or more. Larger mb/ms values can

266

provide higher mechanical energy or a stronger mechanocaloric effect, which may induce an

267

increase in temperature. However, as mb/ms varied from 25 to 100, the change in temperature

268

on the surface of the milled samples was almost the same and the maximum temperature was

269

below 50 °C (Figure 2e), due to the instantaneous temperature produced by ball milling

270

having a short lifetime of 2–10 ns.29 After the ball milling, both PS and Al2O3 were not

271

molten (Figure S3), because the surrounding temperature was much lower than their melting

272

points. In a control test, when the powdery mixtures of PFOA, Al2O3 and PS were heated at

2

3


13


Page 14 of 37

273

100 °C for 2 h, the defluorination efficiency of PFOA was negligible (Figure 2f). Thus, the

274

activation of solid-to-solid reactions by ball milling is not a simple heat activation mechanism.

275

The high energy impact between surfaces of the balls not only produces a transient high

276

temperature but also initiates the rotation, alignment and reorientation of molecules, all of

277

which facilitate changes in the solid surface and chemical bonds. A larger mb/ms can increase

278

the total number of impacts between balls to achieve reaction completion for a given milling

279

time, and any further increase of mb/ms can only act to accelerate to accelerate reaction rate.30,

280

31

(a)

75

50

25

Al2O3-PS Al2O3

(c)

only Al2O3

75

25

25

0

2

4

6

nAl O /nPS 2

(e)

8

10

0

30

100

15

30


100

50

50

25

25

0

15

3

25 50 75 100

(d)

120

75

0

mb/ms


75

100

30

0

0

100

45

T / C

25

120

50

0

PS

50

45

(f)

mb/ms



50

60

Al2O3

0

75

105 100

PFOA-PS

75

PFOA-Al2O3-PS

50

50

25

0

30


14

25

PFOA-Al2O3-PS PFOA-PS

0

120

75

60 Time /min

90

0 120

Residual PS /%

30

75

0

Al2O3-PS

75



100

(b)

PS

0


0

281


100



100

Page 15 of 37


282

Figure 2. Time profiles of degradation (a) and defluorination (b) of PFOA (0.25 g) under 350

283

rpm and mb/ms of 50 with milling agents such as PS (3.95 g), Al2O3 (3.95 g), or the mixture of

284

Al2O3 (2.58 g) and PS (1.37 g) at nAl O /nPS of 5. Effects of nAl O /nPS (c) and mb/ms (d) on

285

degradation (triangles) and defluorination ratios (circles) of PFOA in MC-Al2O3-PS at a 2 h

286

milling. (e) Time profiles of the temperature on the surface of milled samples during the

287

milling. (f) Time profiles of defluorination efficiency of PFOA with PS or the mixture of PS

288

and Al2O3 in solid state by heating to 100 °C.

289

Mineralization of Degradation Intermediates. The degradation intermediates produced

290

during the 1 h MC degradation of PFOA were extracted with ethanol or acetone as described

291

in the Experimental Section. The organic intermediates in the ethanol-extracted solutions

292

were identified by LC/MS, and those in the acetone-extracted solutions were identified by

293

GC/MS.

2

2

3

3

294

When Al2O3 was used as a milling agent, three products were detected in the acetone

295

extracts from milled samples, where CF3(CF2)4CF=CFH was observed as a major product in a

296

yield of 87% with the almost constant selectivity at 88% after a 2 h milling (Figures 3a, S4a,

297

and S5). This indicates that the transformation to CF3(CF2)4CF=CFH was the main

298

degradation pathway of PFOA in MC-Al2O3. When using PS as a milling agent, large

299

amounts of fluoroketones, fluoroethers, and fluoroesters in the acetone-extracted solutions

300

was observed as well as trace amounts of shorter chain PFCAs (such as perfluorohexanoic

301

acid (C5F11COOH, PFHxA) and perfluoroheptanoic acid (C6F13COOH, PFHpA) in

302

ethanol-extracted solutions (Text S1, Figures S4b, S6, and S7). Due to the lack of authentic

303

compounds, we did not quantify the individual intermediates, but measured the total organic

304

carbon (TOC) in the milled sample. As shown in Figure 3b, the TOC resulting from the

305

generated organic products and residual PFOA was 89%, although 61% of the added PFOA


15


306

was degraded after a 2 h milling.

307

When both Al2O3 and PS were used as comilling agents, two components were detected

308

in the ethanol-extracted solutions (Figure S8). The first component was detected at 3.9 min

309

with m/z = 314.9(C5F11COO−), and 272.1(C5F11−) was identified as PFHxA (C5F11COOH),

310

and the second component was detected at 5.6 min with m/z = 362.8(C6F13COO−) and

311

319.7(C6F13−) was identified as PFHpA (C6F13COOH). Four main polyfluorocarbons (PFCs)

312

were detected in the acetone-extracted solutions (Figure S4a). The component that was

313

detected at 1.40 min with the mass fragments of 219.12 (CF3(CF2)3), 169.17 (CF3(CF2)2),

314

119.15 (CF3CF2) and 69.2 (CF3) was assigned to CF3(CF2)2CF2H (Figure S9a). Similarly, the

315

other three components were identified as CF3(CF2)3CF=CF2 at 1.46 min, CF3(CF2)4CF=CFH

316

at 1.50 min, and CF3(CF2)5CF2H at 1.60 min, respectively (Figures S9b-d). The maximum

317

ratio of total organic intermediates was less than 10% of the initial PFOA level, and their

318

accumulation gradually decreased to zero as the ball milling time was prolonged to 2 h

319

(Figure 3c). These comparisons confirm that using both Al2O3 and PS not only alters the

320

pathway of PFOA degradation but also significantly reduces the accumulation of

321

intermediates.

322

In addition, CO2 yields were measured in all three tested mixtures when the gaseous

323

products were collected in NaOH solution. After a 2 h milling, the ratios of CO2 to the initial

324

amount of carbon in PFOA were 18, 7.9, and 98% for MC-Al2O3, MC-PS, and MC-PS-Al2O3,

325

respectively (Figures 3a~c). Throughout the MC treatment, the total carbon included in the

326

residual PFOA, organic intermediates and produced CO2 was almost equal to the initial value

327

for the three tested reactions. Thus, it is emphasized that PFOA was converted to either

328

CF3(CF2)4CF=CFH in MC-Al2O3, or dimeric molecules in MC-PS, while the degradation to

329

CO2 was dominant in MC-PS-Al2O3. Moreover, the TOC removal efficiency of PFOA in

330

MC-PS-Al2O3 was 98%, being 5.4 and 8.9 times higher than those of MC-Al2O3 (18%) and


16

Page 16 of 37

Page 17 of 37

331


MC-PS (11%), respectively. 100

(a)

100

(b)

PFOA

C5F11CF=CFH

50

25

CO2

F2 =C CF

100

0


Carbon distribution /%

0

332

CO2

50 PFHpA PFHxA PFCs

25

0

25

CO2 0 0 100

TC

75

PFOA

50

120

(c) PFOA

75

C3F7CF=CFH

F 1 C5 1

0


75


TC

TOC

TOC removal /%


total carbon (TC)


120

(d)

75 PS Al2O3

50

Al2O3-PS 25

0

120

0


120

333

Figure 3 Distributions and balance of carbon during MC degradation of PFOA in (a)

334

MC-Al2O3, (b) MC-PS and (c) MC-Al2O3-PS. (d) Removal efficiency of total organic carbon

335

(TOC) during MC degradation of PFOA. The measured TOC in (b) includes that in the

336

generated organic products (such as fluoroketones, fluoroethers, and fluoroesters) and residual

337

PFOA.

338

Interaction of PFOA with Comilling Agents and the Transformations. After mixing

339

PFOA with PS and/or Al2O3 by manually grinding in an agate mortar, a small portion of the

340

sample was compressed into a KBr pellet for the FT-IR measurement. Table 1 summarizes

341

the dominant absorption peaks together with their assignments. In FT-IR spectrum of

342

PFOA-PS, a new and weak peak appeared at 1723 cm−1 in addition to the strong absorption


17


343

peak for the carboxyl group of PFOA (Table 1, Figure S10a). Meanwhile, the symmetric

344

stretching (νs) peak of SO3− at 1108 cm−1 significantly decreased in intensity. This indicates

345

that most PFOA molecules exist in the form of carboxylic acid, while a small amount of

346

PFOA is bound to PS through the intermolecular hydrogen bond between the SO3− of PS and

347

the COOH of PFOA.32 For PFOA-Al2O3, the stretching peak of C=O (1768 cm−1) in PFOA

348

disappeared, and νs and the asymmetric stretching peak (νas) of COO− appeared at 1413 and

349

1676 cm−1, respectively, with Δν (= νas−νs) of 263 cm−1. The stretching peaks of the C−F units

350

moved to higher wavenumbers by 4 ~ 7 cm−1 (Table 1, Figure S10b). These changes are

351

related to both the COO−Al coordination in a monodentate mode through the neutralization

352

between COOH and surface OH on Al2O3 (>AlOH) (eq. 4) and hydrogen bond between the F

353

of PFOA and >AlOH (eq. 5).25,33,34 After mixing the three components, the spectral changes

354

related to the carboxyl group and C−F units were similar to those in PFOA-Al2O3 but different

355

from those in PFOA-PS (Table 1, Figure S10c). This indicates that PFOA is bound to the

356

Al2O3 surface rather than that of PS. The deprotonation of PFOA causes the dissociation of

357

hydrogen bonding between PFOA and PS, leading to the recovery of the νs peak of SO3− at

358

1108 cm−1. Unlike the expectation, the weak peak at 1108 cm−1 in PFOA-Al2O3-PS was

359

observed to indicate the presence of other hydrogen bond donors. After mixing PS and Al2O3,

360

the νs peak of SO3− at 1108 cm−1 weakened (Table 1, Figure S10d), implying that the surface

361

OH of Al2O3 and the SO3− of PS act as hydrogen bond donors and acceptors, respectively (eq.

362

6). Thus, it is concluded that Al2O3 anchors the PFOA and PS in ternary mixtures and the two

363

interaction modes including PFOA-Al2O3 and Al2O3-PS do not influence each other. The

364

characteristic absorption peaks of PFOA and PS in MC-Al2O3-PS gradually decreased as the

365

ball milling proceeded. Moreover, the rates of decrease were faster than those for MC-PS

366

(Figure S11), which suggests that the coexistence of Al2O3 promotes the conversion of PFOA

367

and PS.


18

Page 18 of 37

Page 19 of 37


368

 AlOH  C 7 F15 COOH   Al 3 

369

 AlOH   Al 3 

370

 AlOH   OS(O 2 )OOSO 3   AlOH 

371





OOCC 7 F15  H 2 O

OOCC 7 F15   Al 3  





(4)

OOCC 7 F15  HOAl 

OS(O 2 )OOSO 3



(5) (6)

Table 1. Characteristic FT-IR peaks of PFOA, PS, Al2O3, and their mixtures before milling. Samples

Absorption peaka /cm−1 ν(C=O/COO−)

ν(CF)

νas(SO3−)

νs(SO3−)

PFOA

1768

1236, 1204, 1147

−

−

PS

−

−

1303, 1276

1108

Al2O3

−

−

−

−

PFOA-PS

1768, 1722(w)

1236, 1204, 1147

1303, 1276

1108(w)

PFOA-Al2O3

1676, 1413 (w)

1243, 1211, 1151

−

−

PS-Al2O3

−

−

1303, 1276

1108(w)

PFOA-Al2O3-PS

1681, 1415 (w)

1243, 1211, 1151

1303, 1276

1108(w)

372

aThe

373

XPS measurements were conducted to check the elemental changes of Al and S in

374

MC-Al2O3-PS. Before the ball milling, the Al 2p peak approximaterly 74.8 eV was

375

deconvoluted into two peaks at 73.9 and 75.0 eV corresponding to Al−O and Al−OH in Al2O3,

376

respectively (Figure 4a).35 After ball milling for 15 min, the Al−O peak gradually decreased

377

in intensity and a new peak appeared at 76.4 eV, which was assigned to F– bonded to Al

378

(F−Al).36 With an increase in the milling time to 60 min, the Al−O peak became weaker and

379

the intensity of F−Al peak increased. This indicated that more F atoms were eliminated from

380

the PFOA and then bonded to the Al2O3 surface. The sample before milling showed an S 2p

381

peak at 170.3 eV, which was assigned to S2O82− (Figure 4b).37 After ball milling for 1 h, the S

382

2p peak gradually moved to 169.3 eV to be in the proximity of the binding energy of

383

inorganic sulfates.38 This supports the hypothesis the conversion of S2O82– to SO42– took place

384

during the MC degradation of PFOA.

letter “w” in parenthesis represents the weak absorption peak.


19


(a)

(b) Al-OH

Al-O

Al-F

81

Page 20 of 37

78

0 min

0 min

15 min

15 min

60 min

60 min

75 72 Binding energy /eV

385

69

174

172

170 168 Binding energy /eV

166

164

386

Figure 4. High resolution XPS spectra of (a) Al 2p and S 2p for PFOA-Al2O3-PS before and

387

after the ball milling for 15 and 30 min.

388

Identification of Reactive Species. During the ball milling, PS is activated to produce

389

SO4•−,18 and the metal oxide generates oxygen vacancies and free electrons due to a collateral

390

effect of losing lattice oxygen.25, 39, 40 To confirm such reactive species during the milling of

391

Al2O3 and PS, we first employed DPPH• as a probe of free electrons and radicals. When the

392

mixture of DPPH• and Al2O3 is subjected to the ball milling, the absorption peak of DPPH• at

393

525 nm significantly decreased with the appearance of a new peak approximately 425 nm

394

corresponding to DPPH− (Figure 5a),25, 39 which indicates the generation of free electrons on

395

the milled Al2O3 (eq. 7). However, no DPPH− was observed by using PS and Al2O3 as

396

comilling agents, which suggests that the adsorbed PS as a strong oxidant effectively traps the

397

electrons generated on Al2O3 surface (eq. 8). In addition, MC-Al2O3-PS showed a broad

398

absorption peak approximately 440 nm, which is different from that of MC-PS (Figure 5a),

399

but similar to that of the coupling product between DPPH• with •OH.41

400

 Al  O milling   Al  VO (2e - )  0.5O 2

401

e   S2 O8

402

The radical intermediates were also investigated by ESR spectroscopic measurement of

2

surf

 SO 4



surf

 SO 4

2

(7) (8)


20

Page 21 of 37


403

solids recovered from the milled samples in the presence of DMPO as a spin trapping reagent.

404

As shown in Figure 5b, MC-Al2O3 showed no detectable signal in either solid or solution

405

samples. The milled sample of PS (1.0 g) in the presence of DMPO exhibited the strong

406

characteristic peaks of a DMPO-SO4•− adduct with hyperfine splitting constants of aN= 13.2 G,

407

aH= 9.6 and 1.48G (Figure 5b), which confirms the decomposition of PS into SO4•− under the

408

ball milling conditions (eq. 9).42,

409

DMPO-SO4•− signals became weaker due to the decreased amount of PS (Figure 5b).

410

Moreover, an approximately 1:2:2:1 quartet pattern (aH= aN= 14.9 G) was observed, which

411

was assigned to the DMPO-•OH adduct,43 confirming that surface OH group on Al2O3

412

promotes the conversion of SO4•− into •OH (eq. 10). After extracting the milled sample with

413

moisture, formation of the DMPO-•OH adduct was observed in the two mixtures and

414

MC-Al2O3-PS exhibited a stronger DMPO-•OH signal than MC-PS (Figure S12), thereby

415

supporting the conversion of SO4•− to •OH on the Al2O3 surface (•OHsurf). Moreover, the ESR

416

signals for the milled samples of PS-DMPO and Al2O3-PS-DMPO remain considerably

417

intense after the milled samples were placed at room temperature for 22 h and 4 h,

418

respectively (Figure S13). The long-lived DMPO-radical adducts on the milled samples are

419

similar to “environmentally persistent free radicals” that formed on the surface of engineered

420

nanomaterials during combustion or thermal treatments,44 probably because the association of

421

the free radical formed in situ with the surface of milled solids stabilizes the radical. 2

milling   2SO 4

422

S2 O8

423

OH  surf  SO 4

surf



surf



43

For the mixture of Al2O3 (0.7 g) and PS (0.3 g), the

(9)

surf

   OH surf  SO 4

2

(10)


21


(a)

Normalized Abs

1.2

Page 22 of 37

(b)

Al2O3+DPPH Al O +PS+DPPH 2 3

DMPO+Al2O3+PS

0.8 PS+DPPH DMPO+PS

0.4

DPPH DMPO+Al2O3

0.0 300

424

400

500 Wavelength /nm

600

700 3360

3390

3420 3450 Magnetic field /G

3480

3510

425

Figure 5. Normalized UV-vis absorption spectra of DMSO-extracted solutions from the

426

milled samples in the presence of DPPH• (a), where the absorbances of the peaks in the

427

region of 200 ~ 600 nm are normalized to 1.0. ESR spectra of milled samples in the presence

428

of DMPO (b).

429 430

To evaluate the contributions of SO4•−surf and •OHsurf, the effect of surface OH on Al2O3

431

was studied. We preheated Al2O3 powders at 100 °C for 24 h. This preheating decreased the

432

surface water and/or OH− on the Al2O3 by approximately 50% and consequently the

433

defluorination efficiency of PFOA in MC-Al2O3-PS after 2 h milling decreased from 100% to

434

64% (Figure S14). The MC degradation of PFOA was further performed by using 18O-labeled

435

Al2O3 as a comilling agent. Prior to the degradation experiments, Al2O3 powders were

436

dispersed in H218O for 1 h to ensure full oxygen exchange between H218O and the surface OH

437

on Al2O3, following by vacuum drying at 60 °C for 1 h to remove the surface water on Al2O3.

438

For convenience, the obtained

439

the defluorination efficiency of PFOA was reduced from 100% in MC-Al2O3-PS to 70.6% in

440

MC-(18OH-Al2O3)-PS at a 2 h milling (Figure 6a). Meanwhile, the isotope distribution of 18O

441

in another dominant product, CO2, was significantly changed. As shown in Figure 6b, the

442

collected gaseous product shows characteristic FT-IR absorptions of νas(C16O2) at 2361 and

18O-labeled

Al2O3 was named


22

18OH-Al O . 2 3

It was found that

Page 23 of 37


443

2337 cm-1 in MC-PFOA-Al2O3-PS. These two peaks not only shift to lower wavenumbers by

444

approximately 17 cm-1 but also weaken in MC-PFOA-(18OH-Al2O3)-PS at the same milling

445

time, thus indicating that 16O in CO2 exchanged by 18O and the production of CO2 is slower

446

than in MC-PFOA-Al2O3-PS. Accordingly, the mass spectrum of CO2 produced in

447

MC-PFOA-Al2O3-PS exhibits a dominant m/z of 44 (C16O16O, 99.5%), being similar to

448

natural isotope abundance of CO2 in air (99.6%) and obtained from the reaction of Na2CO3

449

with HNO3 (99.3%) (Figures 6c and S15). In contrast, the ratios of CO2 with m/z of 44

450

(C16O16O), 46 (C16O18O) and 48 (C18O18O) in MC-PFOA-(18OH-Al2O3)-PS were changed to

451

63.6%, 32.0% and 4.4%, respectively (Figures 6c). Because the background influence of air

452

may cause a negative deviation in the evaluation of the absolute isotope ratio of

453

( R 18 O/16 O ), we did not calculate R 18 O/16 O . However, the mass area ratio of C16O18O to C18O18O

454

was 7.3, which suggests that one of the two oxygen atoms in CO2 comes from

455

surface OH on Al2O3. After collecting CO2 gas in NaOH solution and converting it to MnCO3

456

precipitant, the stable isotope ratio of

457

(VSMOW) was measured on an elemental analyzer coupled to an isotope ratio mass

458

spectrometer. The δ18O values for the MnCO3 obtained from MC-PFOA-Al2O3-PS and the

459

two reference compounds CO and ref-MnCO3 were 16.08‰, 9.07‰, and 38.02‰,

460

respectively, all of which are close to those of Pee Dee Belemnite, a carbonate isotope

461

standard (30‰ vs. VSMOW, Table S4). In contrast, the δ18O of MnCO3 obtained from

462

MC-PFOA-(18OH-Al2O3)-PS was 3251.50‰, which is 154 times higher than the average

463

value of the abovementioned three controls (21.05‰). The R 18 O/16 O values of CO32− and its

464

precursor CO2 in the case of MC-PFOA-(18OH-Al2O3)-PS were calculated to be 0.32 and 0.61,

465

respectively (Table S4). These results clearly demonstrated that the O atom in the surface OH

466

on Al2O3 can be incorporated into PFOA and finally transformed into CO2. By combining the

467

fact that the conversion of PFOA to CO2 in MC-Al2O3 and MC-PS was less than 20%, it is

18O/16O

18O/16O

18O-labeled

(δ18O, vs. Vienna standard mean ocean water


23


Page 24 of 37

468

reasonable to propose that the surface OH− on Al2O3 is oxidized by SO4•−surf to •OHsurf, which

469

is then predominant species in the mineralization of PFOA.


100

(b)

(a)

15 min 60 min

75

900

18

MC-( OH-Al2O3)-PS

50

15 min

18

( OH-Al2O3)-PS Al2O3-PS

25

0

60 min MC-Al2O3-PS

0

60


240

2500

2400 2300 -1 Wavenumber /cm

2200

(c) 8.0x10

6

CO2 obtained from

Intensity

Na2CO3+HNO3

6.0x10

6

4.0x10

6

2.0x10

6

MC-PFOA-Al2O3-PS 18

MC-PFOA-( OH-Al2O3)-PS Air

0.0 38

40

42

44

46

m/z

48

50

52

54

470 471

Figure 6. (a) Time profiles of defluorination efficiency of PFOA in MC-Al2O3-PS and

472

MC-PFOA-(18OH-Al2O3)-PS. (b) FT-IR and (c) mass spectra of collected gaseous products

473

from MC degradation of PFOA. (d) Isotopic oxygen composition (δ18O) of MnCO3 obtained

474

from degradation products of PFOA in MC-Al2O3-PS at a 2 h milling (S1-MnCO3) and

475

MC-(18OH-Al2O3)-PS at a 4 h milling (S2-MnCO3). CO2 obtained from the reaction Na2CO3

476

with HNO3 and indoor air, commercial CO gas and MnCO3 obtained from the reaction

477

Na2CO3 with MnSO4 (ref-MnCO3) were given as references in (c) and (d).

478

Reaction Mechanism. Based on the above results, the reaction mechanism of degradation of

479

PFOA using both Al2O3 and PS under the conditions of ball milling is proposed in Scheme 1.

480

After mixing PFOA with one agent, PFOA was tightly bound to Al2O3 through the COO−Al


24

Page 25 of 37


481

bond and >AlOHF−C hydrogen bonding (eqs. 4 and 5), while the O2− of PS acted as the

482

proton accepter to bind PFOA through the hydrogen bond in MC-PS. In the ternary mixture,

483

the interaction between PFOA and Al2O3 does not change, but Al2O3 replaces PFOA to bind

484

PS via hydrogen bonding of >AlOHO2− (PS) (eq. 6). That is, Al2O3 strongly anchors the

485

PFOA and PS (Scheme 1). During the MC treatment, the high-energy impact has multiple

486

roles in the activation of solid-to-solid reactions by producing a transient temperature and

487

initiating the rotation, alignment and reorientation of molecules. On one hand, ball milling

488

induces the decarboxylation reaction of PFOA, thereby yielding CO2 and an unstable

489

perfluoroheptyl anion (−C7F15) on the Al2O3 surface (eq. 11) which preferentially reacts with

490

the proton of an OH group on Al2O3 to yield C7F15H (eq. 12). On the other hand, the

491

high-energy impact facilitates the decomposition of PS into SO4•−, which consequently

492

oxidizes OH−surf on Al2O3 to •OHsurf. As demonstrated in Figures 5 and 6, the active species

493

for the degradation of PFOA is •OHsurf in MC-Al2O3-PS. Hydrogen abstraction from C7F15H

494

by •OHsurf is followed by the formation of •C7F15 (eq. 13), which couples with •OHsurf to yield

495

C7F15OH (eq. 14). C7F15OH transforms into C6F13COF through the elimination of HF (eq. 15)

496

and reacts with the surface water to yield C6F13COOH together with H+ and F− (eq. 16).

497

Similarly, C6F13COOH is bound on the Al2O3 surface via the COO−Al and AlF

498

interactions, which activates the C−COO bond and C−F bonds, respectively (Scheme 1). This

499

leads the decarboxylation, which yields –C6F13, to take place as a second reaction on the

500

Al2O3 surface.

501

C 7 F15 COO   Al  milling   C 7 F15  Al   CO 2

(11)

502

(  C 7 F15 )

(12)

503

(C 7 F15 H) surf   OH surf  (  C 7 F15 ) surf  H 2 O

(13)

504

(  C 7 F15 ) surf   OH surf  (C 7 F15 OH) surf

(14)

surf

 H  surf  (C 7 F15 H) surf


25


Page 26 of 37

505

(C 7 F15 OH) surf  (C 6 F13COF) surf  (HF (H   F  )) surf

(15)

506

C 6 F13COFsurf  H 2 O surf  C 6 F13COOH surf  (HF (H   F  )) surf

(16)

507 508

Scheme 1. A schematic representation for MC degradation of PFOA in the presence of both

509

Al2O3 and PS.

510 511

The

successive

degradation

including

decarboxylation,

protonation,

512

hydrogen-abstraction, radical coupling, elimination of HF and hydrolysis (eqs. 11 ~ 16), occur

513

to yield short-chain PFCA products until CF3COOH is reached, which degrades to −CF3 on

514

the Al2O3 surface (eq. 11). Degradation and defluorination are repeated to yield CO2, H+, and

515

F−as the final products (eqs. 12 ~ 15). Thus, complete degradation and defluorination are

516

achieved. It is noted that the mineralization of 1 mole PFOA consumes 14 moles •OHsurf, of

517

which 7 moles •OHsurf are converted to CO2, and the others to H2O (Scheme S1). Moreover,

518

the mineralization of 1 mole PFOA generates 8 moles CO2, of which 7 moles come from the

519

produced short-chain PFCAs (Scheme S1). Each short-chain PFCA molecule (CnF2n+1COOH)

520

contains two external oxygen atoms: one from the addition of •OHsurf (eqs. 14 and 15, Scheme

521

S1), and the other from surface water via the hydrolysis reaction (eq. 16, Scheme S1). In

522

18O-labeling

experiments, all the surface OH− groups on Al2O3 were exchanged by the 18OH


26

Page 27 of 37


523

of H218O, followed by drying at 60 °C for 1 h to remove surface water on the Al2O3. During

524

the MC treatment, the oxidation of surface 18OH− on Al2O3 by SO4•−surf yields •18OHsurf, while

525

surface water comes from moisture at natural isotope abundance (H216O), giving a theoretical

526

R 18 O/16 O in the produced CO2 of 0.78 as shown in eq. 17. The expected observations of

527

C16O18O as a main product and an

528

demonstrate that the MC degradation of PFOA in MC-Al2O3-PS follows a •OHsurf involved

529

pathway (eqs. 11 ~ 16).

R 18 O/16 O (CO2) = 0.6 (Figures 6b and 6c) clearly

C7F15COOHsurf + 718OH−surf + 7H+surf + 14•18OHsurf + 7H216Osurf 530

→ CO2 + 7C16O18O + 7H218O + 7H218O + 15HFsurf

(17)

531

From eq. 17, a complete defluorination of 1 mole PFOA may yield 15 moles F− and consumes

532

14 moles •OHsurf, being equivalent to 7 moles PS as calculated from eqs. 9 and 10. According

533

to eq. 3, the utilization efficiency of PS (PS) was calculated from the ratio of the

534

stoichiometrically required PS for generating F− to the actual consumption. For the MC

535

degradation of PFOA (0.25 g) at nAl O /nPS /nPFOA of 42: 8.4: 1, the residual PS decreased to

536

3.7% when the defluorination efficiency of PFOA was 100% (Figure S16a). The value of PS

537

in MC-Al2O3-PS remained at 87% during the defluorination of PFOA (Figure S16b). This

538

value is much higher than those for the defluorination of PFOA in aqueous solution (8% ~

539

22%) and aqueous soil slurry (< 1%).45−48

2

3

540

It should be noted that the main active species responsible for the degradation of PFOA

541

are the “free” electrons in MC-Al2O3 and SO4•− in MC-PS, respectively. This results in

542

different degradation pathways for PFOA from •OHsurf involved reactions in MC-Al2O3-PS.

543

In MC-Al2O3, the primary step of degradation of PFOA is the decarboxylation reaction (eq.

544

11) and then the “free” electrons on Al2O3 (eq. 7) enable the conversion of the perfluoroalkyl

545

anion to C5F11CF=CFH through protonation and reductive defluorination.25 In MC-PS, SO4•−


27


546

oxidizes the PFOA adsorbed on the surface of PS powders to form unstable •C7F15 through a

547

one-electron oxidation, deprotonation, and decarboxylation (Text S2, eqs. S1 ~ 3). Further

548

reactions between •C7F15 and PFOA yield C7F15COOC7F15, C7F15COC7F15, C7F15OC7F15, and

549

C7F15COC6F13 as the main products (Text S2, eqs. S4 ~ 7, Figure S7). Thus, the TOC removal

550

and defluorination ratios were as low as 11% and 6.4%, respectively (Figures 1b and 2b),

551

although 61% of the PFOA was degraded in MC-PS after a 2 h milling. It is also noted that

552

the defluorination efficiency of PFOA in MC-PS (< 10%) is lower than those (> 70%)

553

induced by photolysis or thermolysis of PS in aqueous solution,45−47 due to the different

554

degradation pathways of PFOA. Upon treatment with PS in water, SO4•− can be converted

555

into •OH, and both of them contribute to the degradation of PFOA. The unstable •C7F15

556

generated from oxidation by SO4•− usually reacts with •OH, dissolved O2 and/or sufficient

557

H2O to yield C6F13COOH, which follows an analogous mechanism to produce short-chain

558

PFCAs products until complete mineralization and defluorination occurs.45−48 The fast

559

diffusion of chemical species in water will not cause PFCA molecules to remain close to the

560

PS species until the parent PFCA molecule is completely defluorinated. As a result, quite

561

large amounts of short-chain PFCAs were detected as major products during the

562

defluorination of PFOA in PS-involved aqueous solution.45−48

563

Applications and Implications. We found complete degradation and defluorination of

564

PFOA through activation of C−COO− and C−F bonds and generation of •OH by the MC

565

treatment using Al2O3 and PS as comilling agents, respectively. The MC treatment was also

566

used to degrade the homologues of PFOA such as PFPA, PFHxA, and PFHpA under the same

567

conditions (i.e., mAl O /mPS /mPFCA , 2.58: 1.37: 0.25; mb/ms, 50:1). After ball milling for 2 h,

568

such PFOA homologues are degraded with a defluorination efficiency of 100% (Figure S17).

569

Perfluorooctane sulfonic acid (PFOS) is another important perfluorinated pollutant. In general,

570

PFOS is much more resistant to oxidative degradation by SO4•− than PFCAs, and hence, no

2

3


28

Page 28 of 37

Page 29 of 37


571

transformation of PFOS was observed in aqueous solution or soil slurry in the presence of

572

heat-activated PS.47,48 We used the developed MC method to degrade PFOS, and found that

573

56% of the added PFOS was slowly degraded in MC-Al2O3-PS with a defluorination

574

efficiency of 9.7% after a 2 h milling (Figure S18). The considerably slower degradation of

575

PFOS than PFOA was in good agreement with the abovementioned opinion that PFOS is

576

much more resistant to oxidative degradation by SO4•−. Furthermore, the enhanced

577

degradation of PFOS in MC-Al2O3-PS relative to that in MC-PS may also support the

578

important role of the Al2O3 surface in the activation of the perfluorinated molecules. The

579

lower extent of defluorination of PFOS in MC-Al2O3-PS was due to the C−SO3− bond in

580

PFOS being cleaved by the oxidation to produce SO3 and •C7F15 (eq. 18) with the generated

581

SO3 reacting with OH−surf (eq. 19), which suppressed the transformation of SO4•− to •OHsurf

582

(eq. 10), significantly decreasing the generation of •OH for its coupling with •C7F15. Finally,

583

the main fortune of •C7F15 is to form dimers with itself without considerable elimination of F

584

atoms. 



   C 7 F15  SO 3  SO 4

585

C 7 F15SO 3  SO 4

586

SO3  OH  surf   SO 4  H 

surf

2

2

(18) (19)

587

The advanced oxidation processes (AOPs) using •OH or SO4•− have drawn interest in the

588

degradation of organic pollutants.8−12 However, there are limited AOPs capable of the

589

complete defluorination and mineralization of PFOA pollutants in solids. The present study of

590

MC treatment using both Al2O3 and PS provides valuable information for the design of AOPs

591

for the complete defluorination and mineralization of refractory PFCAs. Moreover, the

592

normalized energy consumption for releasing one micro-molar fluorine ions from the

593

degradation of PFOA (E, kJ μmol−1) was evaluated to be approximately 0.15 kJ μmol−1 in

594

MC-Al2O3-PS. This is the same order of magnitude in previously reported MC treatments,21

595

but higher energy efficiency is obtained using this method than using other existing methods


29


Page 30 of 37

596

(Table 2). It was noted that the degradation and defluorination efficiencies of PFOS in

597

MC-Al2O3-PS were higher than those (44% and 3.4%, respectively) in MC-PS, but lower than

598

those in MC-Al2O3 (96% and 21%) (Figure S18). This inspires us to develop in the future an

599

appropriate MC treatment method to degrade PFOS.

600 601

Table 2. Comparison of energy consumption for the degradation of PFOA by different

602

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

Deflurination

Energy

(W)

efficiency (%)

(kJ mol−1)a

750

100 (2 h)

0.15 b

750

97 (4 h)

0.39 b

21

750

19 (4 h)

1.99 b

21

23

33.7 (4 h)

1.64

49

10

59.3 (24 h)

24.3

50

200

59.1 (4 h)

10.8

45

9

46 (24 h)

28.2

51

> 50 c

81 (5 h)

> 5.10

52

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 L 0.5 g L−1 In2O3 0.02 mM PFOA, 0.2 L, 0.5 g L-1 BiOCl 29.6 mol 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

Ref. This work

603

a

Energy consumed to release one micro-molar F− ions from the degradation PFOA, which is

604

calculated from eq. S11 (Text S3). b Planetary ball mill is a robust floor model with 4 grinding

605

stations, and thus the treated PFOA should be 4 times of the added pollutants.

606

power includes the ultrasonic irradiation (50 W) and the employed power in electrochemical

607

experiment, but the latter of them is unknown.


30

c

The total

Page 31 of 37


608

ASSOCIATED CONTENT

609

Supporting Information

610

Supporting texts (Text S1−S3), figures (Figure S1−S18), and table (Table S1−S4), as

611

mentioned in the supporting information. This material is available free of charge via the

612

Internet at http://pubs.acs.org.

613 614



615

Corresponding Author

616

** Prof. Lihua Zhu, Fax: +86 27 8754 3632; E-mail:[email protected].

617

Prof. HeqingTang, Fax: +86 27 6784 3990; E-mail: [email protected].

618

Notes

619

The authors declare no competing financial interest.

AUTHOR INFORMATION

620 621



622

This research was financially supported by the National Natural Science Foundation of China

623

(Grants Nos. 21777194 and 21477043). We thank to Prof. Anmin Zhen and, Dr. Shenhui Li,

624

Wuhan Institute of Physics and Mathematics of the Chinese Academy of Science, and Dr.

625

Zhiwu Yu, High Magnetic Field Laboratory of the Chinese Academy of Science for their

626

help.

ACKNOWLEDGEMENTS

627 628



629

(1) Zareitalabad, P.; Siemens, J.; Hamer, M.; Amelung, W. Perfluorooctanoic acid (PFOA)

630

and perfluorooctanesulfonic acid (PFOS) in surface waters, sediments, soils and

631

wastewater-A review on concentrations and distribution coefficients. Chemosphere 2013, 91,

632

725−732.

633

(2) Houde, M.; De Silva, A. O.; Muir, D. C.; Letcher, R. J. Monitoring of perfluorinated

634

compounds in aquatic biota: an updated review. Environ. Sci. Technol. 2011, 45, 7962−7973.

REFERENCES


31


635

(3) Xiao, F.; Simcik, M. F.; Halbach, T. R.; Gulliver, J. S. Perfluorooctane sulfonate (PFOS)

636

and perfluorooctanoate (PFOA) in soils and groundwater of a U.S. metropolitan area:

637

migration and implications for human exposure. Water Res. 2015, 72, 64−74.

638

(4) Cagnetta, G.; Robertson, J.; Huang, J.; Zhang, K.; Yu, G. Mechanochemical destruction of

639

halogenated organic pollutants: A critical review. J. Hazard. Mater. 2016, 313, 85−102.

640

(5) Song, Z.; Tang, H.; Wang, N.; Zhu, L. Reductive defluorination of perfluorooctanoic acid

641

by hydrated electrons in a sulfite-mediated UV photochemical system. J. Hazard. Mater.

642

2013, 262, 332−338.

643

(6) Gu, Y.; Dong, W.; Luo, C.; Liu, T. Efficient reductive decomposition of perfluorooctane

644

sulfonate in a high photon flux UV/sulfite system. Environ. Sci. Technol. 2016, 50,

645

10554−10561.

646

(7) Tang, H.; Xiang, Q.; Lei, M.; Yan, J.; Zhu, L.; Zou, J. Efficient degradation of

647

perfluorooctanoic acid by UV-Fenton process. Chem. Eng. J. 2012, 184, 156−162.

648

(8) Liu, Y.; Chen, S.; Quan, X.; Yu, H.; Zhao, H.; Zhang, Y. Efficient mineralization of

649

perfluorooctanoate by electro-Fenton with H2O2 electro-generated on hierarchically porous

650

carbon. Environ. Sci. Technol. 2015, 49, 13528−13533.

651

(9) Song, C.; Chen, P.; Wang, C.; Zhu, L. Photodegradation of perfluorooctanoic acid by

652

synthesized TiO2-MWCNT composites under 365nm UV irradiation. Chemosphere 2012, 86,

653

853−859.

654

(10) Sun, B.; Ma,J.; Sedlak, D. L. Chemisorption of perfluorooctanoic acid on powdered

655

activated carbon initiated by persulfate in aqueous solution. Environ. Sci. Technol. 2016, 50,

656

7618−7624.

657

(11) da Silva-Rackov, C. K. O.; Lawal, W. A.; Nfodzo, P. A.; Vianna, M. M. G. R.; do

658

Nascimento, C. A. O.; Choi, H. Degradation of PFOA by hydrogen peroxide and persulfate

659

activated by iron-modified diatomite. Appl. Catal. B Environ. 2016, 192, 253−259.


32

Page 32 of 37

Page 33 of 37


660

(12) Qian, Y.; Guo, X.; Zhang, Y.; Peng, Y.; Sun, P.; Huang, C. H.; Niu, J.; Zhou, X.;

661

Crittenden, J. C. Perfluorooctanoic acid degradation using UV-persulfate process: Modeling

662

of the degradation and chlorate formation. Environ. Sci. Technol. 2016, 50, 772−781.

663

(13) Wang, F.; Lu, X.; Li, X. Y.; Shih, K. Effectiveness and mechanisms of defluorination of

664

perfluorinated alkyl substances by calcium compounds during waste thermal treatment.

665

Environ. Sci. Technol. 2015, 49, 5672−5680.

666

(14) Zhang, Q.; Matsumoto, H.; Saito, F.; Baron, M. Debromination of hexabromobenzene by

667

its co-grinding with CaO. Chemosphere 2002, 48, 787−793.

668

(15) Zhang, Q.; Saito, F.; Ikoma, T.; Tero-Kubota, S. Effects of quartz addition on the

669

mechanochemical dechlorination of chlorobiphenyl by using CaO. Environ. Sci. Technol.

670

2001, 35, 4933−4935.

671

(16) Zhang, W.; Huang, J.; Xu, F.; Deng, S.; Zhu, W.; Yu, G. Mechanochemical destruction

672

of pentachloronitrobenzene with reactive iron powder. J. Hazard. Mater. 2011, 198, 275−281.

673

(17) Zhang, K.; Huang, J.; Zhang, W.; Yu, Y.; Deng, S.; Yu, G. Mechanochemical

674

degradation of tetrabromobisphenol A: Performance, products and pathway. J. Hazard.

675

Mater. 2012, 243, 278−285.

676

(18) Huang, A.; Zhang, Z.; Wang, N.; Zhu, L.; Zou, J. Green mechanochemical oxidative

677

decomposition of powdery decabromodiphenyl ether with persulfate. J. Hazard. Mater. 2016,

678

302, 158−165.

679

(19) Liu, X.; Zhang, X.; Zhang, K.; Qi, C. Sodium persulfate-assisted mechanochemical

680

degradation of tetrabromobisphenol A: Efficacy, products and pathway. Chemosphere 2016,

681

150, 551−558.

682

(20) Yan, X.; Liu, X.; Qi, C.; Wang, D.; Lin, C. Mechanochemical destruction of a

683

chlorinated polyfluorinated ether sulfonate (F-53B, a PFOS alternative) assisted by sodium

684

persulfate. RSC Adv. 2015, 5, 85785−85790.


33


685

(21) Zhang, K.; Huang, J.; Yu, G.; Zhang, Q.; Deng, S.; Wang, B. Destruction of

686

perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) by ball milling.

687

Environ. Sci. Technol. 2013, 47, 6471−6477.

688

(22) Zhang, K.; Cao, Z.; Huang, J.; Deng, S.; Wang, B.; Yu, G. Mechanochemical destruction

689

of Chinese PFOS alternative F-53B. Chem. Eng. J. 2016, 286, 387–393.

690

(23) Chen, W.; Hooper, T.; Ng, J.; White, A. J. P.; Crimmin, M. R. Palladium-catalyzed

691

carbon-fluorine and carbon-hydrogen bond alumination of fluoroarenes and heteroarenes.

692

Angew. Chem. Int. Ed. 2017, 56, 12687−12691.

693

(24) Senaweera, S. M.; Singh, A.; Weaver, J. D. Photocatalytic hydrodefluorination: Facile

694

access to partially fluorinated aromatics. J. Am. Chem. Soc. 2014, 136, 3002−3005.

695

(25) Lv, H.; Wang, N.; Zhu, L.; Zhou, Y.; Li, W.; Tang, H. Alumina-mediated

696

mechanochemical method for simultaneously degrading perfluorooctanoic acid and

697

synthesizing a polyfluoroalkene. Green Chem. 2018, 20, 2526−2533.

698

(26) Cagnetta, G.; Zhang, Q., Huang, J.; Lu, M.; Wang, B.; Wang, Y.; Deng, S.; Yu, G.

699

Mechanochemical destruction of perfluorinated pollutants andmechanosynthesis of lanthanum

700

oxyfluoride: A waste-to-materialsprocess. Chem. Eng. J. 2017, 316, 1078−1090.

701

(27) Zhou, Q.; Chen, M.; Zhu, L.; Tang, H. Determination of perfluorinated carboxylic acids

702

in water using liquid chromatography coupled to a corona-charged aerosol detector. Talanta

703

2015, 136, 35−41.

704

(28) Liang, C. J.; Huang, C. F.; Mohanty, N.; Kurakalva, R.M. A rapid spectrophotometric

705

determination of persulfate anion in ISCO. Chemosphere 2008, 73, 1540−1543.

706

(29) Delogu, F. Molecular dynamics of atomic rearrangements at perturbed surfaces. Phys.

707

Rev. B 2009, 80, 014115-1−014115-8

708

(30) G. Cagnetta, K. Zhang, Q. Zhang, J. Huang, G. Yu. Mechanochemical pre-treatment for

709

viable recycling of plastic waste containing haloorganics. Waste Manage. 2018, 75, 181–186.


34

Page 34 of 37

Page 35 of 37


710

(31) Cagnetta, G., Huang, J., Wang, B., Deng, S., Yu, G. A comprehensive kinetic model for

711

mechanochemical destruction of persistent organic pollutants. Chem. Eng. J. 2016, 291, 30–

712

38.

713

(32) Liu, Q.; Huang, A.; Wang, N.; Zheng, G.; Zhu, L. Rapid fluorometric determination of

714

perfluorooctanoic acid by its quenching effect on the fluorescence of quantum dots. J. Lumin.

715

2015, 161, 374–381.

716

(33) Nara, M.; Torii, H.; Tasumi, M. Correlation between the vibrational frequencies of the

717

carboxylate group and the types of its coordination to a metal ion: An ab Initiomolecular

718

orbital study. J. Phys. Chem. 1996, 100, 19812−19817.

719

(34) Wang, M.; Wang, N.; Tang, H.; Cao, M.; She, Y.; Zhu, L. Surface modification of

720

nano-Fe3O4 with EDTA and its use in H2O2 activation for removing organic pollutants. Catal.

721

Sci. Technol. 2012, 2, 187−194.

722

(35) Lu, W.; Iwasa, Y.; Ou, Y.; Jinno, D.; Kamiyama, S.; Petersen, P. M.; Ou, H. Effective

723

optimization of surface passivation on porous silicon carbide using atomic layer deposited

724

Al2O3. RSC Adv. 2017, 7, 8090−8097.

725

(36) Shimizu, K.; Driver, G. W.; Lucas, M.; Sparrman, T.; Shchukarev, A.; Boily, J. F.

726

Bifluoride ([HF2]−) formation at the fluoridated aluminium hydroxide/water interface. Dalton

727

Trans. 2016, 45, 9045−9050.

728

(37) Groveman, S.; Peng, J.; Itin, B.; Diallo, I.; Pratt, L. M.; Greer, A.; Biddinger, E. J.;

729

Greenbaum, S. G.; Drain, C. M.; Francesconi, L.; Vittadello, M. The role of ozone in the

730

formation and structural evolution of graphene oxide obtained from nanographite. Carbon

731

2017, 122, 411−421.

732

(38) Pany, S.; Naik, B.; Martha, S.; Parida, K. Plasmon induced nano Au particle decorated

733

over S,N-modified TiO2 for exceptional photocatalytic hydrogen evolution under visible light.

734

ACS Appl. Mater. Interfaces 2014, 6, 839−846.


35


Page 36 of 37

735

(39) Cagnetta, G.; Huang, J.; Lu, M.; Wang, B.; Wang, Y.; Deng, S.; Yu,G. Defect

736

engineered oxides for enhanced mechanochemical destruction of halogenated organic

737

pollutants. Chemosphere 2017, 184, 879−883.

738

(40) Chai, H.; Zhang, Z.; Zhou, Y.; Zhu, L.; Lv, H.; Wang, N. Roles of intrinsic Mn3+ sites

739

and

740

decabromodiphenyl ether with manganese dioxide. Chemosphere 2018, 207, 41−49.

741

(41) Hristea, E. N.; Hillebrand, M.; Caproiu, M. T.; Caldararu, H.; Constantinescu, T.;

742

Balaban, A. T. Scavenging the hydroxyl radical by 2,2-diphenyl-1-picrylhydrazyl. ARKIVOC

743

2002, ii, 123−132.

744

(42) Huang, Z.; Bao, H.; Yao, Y.; Lu, W.; Chen, W. Novel green activation processes and

745

mechanism of peroxymonosulfate based on supported cobalt phthalocyanine catalyst. Appl.

746

Catal. B Environ. 2014, 154−155, 36−43.

747

(43) Fang, G.; Gao, J.; Dionysiou, D. D.; Liu, C.; Zhou, D. Activation of persulfate by

748

quinones: Free radical reactions and implication for the degradation of PCBs. Environ. Sci.

749

Technol. 2013, 47, 4605−4611.

750

(44) Vejerano, E. P.; Rao, G.; Khachatryan, L.; Cormier, S. A.; Lomnicki, S. Environmentally

751

persistent free radicals: Insights on a new class of pollutants. Environ. Sci. Technol. 2018, 52,

752

2468−2481.

753

(45) Hori, H., Yamamoto, A., Hayakawa, E., Taniyasu, S., Yamashita, N., Kutsuna, S.,

754

Kiatagawa, H., Arakawa, R. Efficient decomposition of environmentally persistent

755

perfluorocarboxylic acids by use of persulfate as a photochemical oxidant. Environ. Sci.

756

Technol. 2005, 39, 2383−2388.

757

(46) Hori, H.; Nagaoka, Y.; Murayama, M.; Kutsuna, S. Efficient decomposition of

758

perfluorocarboxylic acids and alternative fluorochemical surfactants in hot water. Environ.

759

Sci. Technol. 2008, 42, 7438−7443.

lattice

oxygen

in

mechanochemical

debromination


36

and

mineralization

of

Page 37 of 37


760

(47) Bruton, T. A.; Sedlak, D. L. Treatment of perfluoroalkyl acids by heat-activated

761

persulfate under conditions representative of in situ chemical oxidation. Chemosphere 2018,

762

206, 457−464.

763

(48) Park, S.; Lee, L. S.; Medina, V. F.; Zull, A.; Waisner, S. Heat-activated persulfate

764

oxidation of PFOA, 6:2 fluorotelomer sulfonate, and PFOS under conditions suitable for

765

in-situ groundwater remediation. Chemosphere 2016, 145, 376−383.

766

(49) Li, X.; Zhang, P.; Jin, L.; Shao, T.; Li, Z.; Cao, J. Efficient photocatalytic decomposition

767

of perfluorooctanoic acid by indium oxide and its mechanism. Environ. Sci. Technol. 2012,

768

46, 5528−5534.

769

(50) Song, Z.; Dong, X.; Wang, N.; Zhu, L.; Luo, Z.; Fang, J.; Xiong, C. Efficient

770

photocatalytic defluorination of perfluorooctanoic acid over BiOCl nanosheets via a hole

771

direct oxidation mechanism, Chem. Eng. J. 2017, 317, 925−934.

772

(51) Tang, H.; Xiang, Q.; Lei M.; Yan, J.; Zhu, L.; Zou, J. Efficient degradation of

773

perfluorooctanoic acid by UV-Fenton process, Chem. Eng. J. 2012, 184, 156–162.

774

(52) Zhao, H.; Gao, J.; Zhao, G.; Fan, J.; Wang, Y.; Wang, Y. Fabrication of novel

775

SnO2-Sb/carbon

776

perfluorooctanoate with high catalytic efficiency. Appl. Catal. B Environ. 2013, 136,

777

278−286.

aerogel

electrode

for

ultrasonic


37

electrochemical

oxidation

of

Complete defluorination and mineralization of perfluorooctanoic acid

Recommend Documents