Reductive defluorination of perfluorooctanoic acid by hydrated electrons in a sulfite-mediated UV photochemical system

Journal of Hazardous Materials 262 (2013) 332–338

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Reductive defluorination of perfluorooctanoic acid by hydrated electrons in a sulfite-mediated UV photochemical system Zhou Song a,b , Heqing Tang a,∗ , Nan Wang b,∗∗ , Lihua Zhu b a Key Laboratory of Catalysis and Materials Science of the State Ethnic Affairs Commission and Ministry of Education, College of Chemistry and Materials Science, South-Central University for Nationalities, Wuhan 430074, PR China b College of Chemistry and Chemical Engineering, Huazhong University of Science and Technology, Wuhan 430074, PR China

h i g h l i g h t s

g r a p h i c a l

a b s t r a c t

• A new reductive method for PFOA

• • •

a r t i c l e

i n f o

Article history: Received 26 June 2013 Received in revised form 22 August 2013 Accepted 23 August 2013 Available online 31 August 2013 Keywords: Perfluorooctanoic acid Photochemical reduction Hydrated electron Sulfite

C5 F11

C6F13 C COOH

Defluorination C COOH

F

F

eaq

100

F

F

-F-

H C6F13 C COOH F

-CH 2 hv -F-

eaq

H C 6F13 C COOH

Defluorination efficiency / %

•

defluorination was established by sulfite-mediated photolysis. The defluorination of PFOA was dependent on sulfite concentration and solution pH. A defluorination ratio of PFOA as high as 88.5% was achieved after reaction of 24 h. A few of perfluorinated sulfonates were detected as intermediates during the degradation of PFOA. A mechanism was proposed for the reductive defluorination of PFOA by hydrated electrons.

80 60 40 20

H

0

0

8

16

24

Time / h

a b s t r a c t A method for reductive degradation of perfluorooctanoic acid (PFOA) was established by using a sulfite/UV process. This process led to a PFOA removal of 100% at about 1 h and a defluorination ratio of 88.5% at reaction time of 24 h under N2 atmosphere, whereas the use of either UV irradiation or SO3 2− alone induced little defluorination of PFOA under the same conditions. It was confirmed that the reductive defluorination of PFOA was achieved by hydrated electrons being generated from the photo-conversion of SO3 2− as a mediator. Theoretical reaction kinetic analysis demonstrated that the generation of hydrated electrons was promoted by increasing either SO3 2− concentration or solution pH, leading to the acceleration of the PFOA defluorination. Accompanying the reduction of PFOA, a small amount of short-chain perfluorocarboxylic acids, less fluorinated carboxylic acids and perfluorinated alkyl sulfonates were generated, all of which were able to be further degraded with further releasing of fluoride ions. Based on the generation, accumulation and distribution of intermediates, hydrated electrons induced defluorination pathway of PFOA was proposed in a sulfite-mediated UV photochemical system. © 2013 Elsevier B.V. All rights reserved.

1. Introduction

∗ Corresponding author. Tel.: +86 27 67843323; fax: +86 27 67843323. ∗∗ Corresponding author. Tel.: +86 27 87543632; fax: +86 27 87543632. E-mail addresses: [email protected], [email protected] (H. Tang), [email protected] (N. Wang). 0304-3894/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.jhazmat.2013.08.059

Perfluorooctanoic acid (C7 F15 COOH, PFOA), as a main endproduct during most perfluorinated compounds (PFCs) degradation through natural processes [1], has been recognized as a persistent organic pollutant and hence received much attention [2]. PFOA and its perfluorinated precursors are widely used as fire retardants, photolithographic film, carpet cleaners and paper coatings due to

Z. Song et al. / Journal of Hazardous Materials 262 (2013) 332–338

their unique hydrophobic and oleophobic properties [3–5]. As the use of PFCs has increased, some of them (typically, PFOA) have been detected in the environment across the globe [6,7]. Therefore, much effort has been devoted to eliminate its adverse effect on human health and ecosystem [3,7]. Several oxidative processes including direct photolysis [8,9], photochemical oxidation [9–12], and photocatalytic oxidation [13–17] have been employed to degrade PFOA. Although it was reported that photogenerated holes in In2 O3 were efficient for decomposing PFOA, many of the reported processes showed slow defluorination rates of PFOA, probably because the electronegative fluorine substituents envelope completely the carbon skeleton and shield it from the chemical attack of • OH. In contrast, the strong electronegativity of fluorine atom(s) may act as the reductive reaction center(s) for defluorination. Moreover, since the toxicity and persistent of PFOA is closely related to the fluorine atoms, the cleavage of C F bond is the critical step for eliminating its adverse effect. Thus, a reduction process may be a better alternative for the removal of PFOA. Hydrated electron (e− aq ) is a powerful reducing agent with a standard reduction potential of −2.9 V, and it can react rapidly with halogenated organic compounds [18]. Huang et al. [19] demonstrated the feasibility of the reaction between e− aq and PFOA by using the laser flush photolysis of K4 Fe(CN)6 in aqueous solution, and estimated the second-order rate constant to be 1.7 × 107 M−1 s−1 for the reaction of e− aq with PFOA. Qu et al. [20] and Park et al. [21] investigated the e− aq -induced photo-reductive defluorination of PFOA using KI as a mediator. However, these methods suffer from some problems such as the limited applications of laser flush photolysis in the practical treatment of massive PFOA and the potential insalubrious effect of iodides on humans [22]. Thus, it is required to find more economical, eco-friendly and massively operable reductive processes for the PFOA defluorination. In the present work, we are focused on SO3 2− -mediated photolysis, because it may function as an alternative source of e− aq . Fischer and Warneck reported that the photooxidation of hydrogen sulfite in aqueous solution could generate e− aq for hydrogen production [23]. Li et al. used a sulfite/UV process for the reductive dechlorination of monochloroacetic acid [24]. In this work, we firstly employed the sulfite/UV process to reduce PFOA, and achieved a defluorination ratio as high as 88.5% at reaction time of 24 h. Along with the investigation of the important reaction parameters, the mechanism of the PFOA degradation was also explored in detail.

2. Materials and methods 2.1. Materials PFOA (C7 F15 COOH, 96%), heptafluorobutyric acid (PFBA, C3 F7 COOH, 99%) and pentafluoropropionic acid (PFPrA, C2 F5 COOH, 97%) were purchased from Acros (New Jersey, USA). Perfluoroheptanoic acid (PFHpA, C6 F13 COOH, 98%) was purchased from Alfa Aesar (Lancs, UK). Undecafluorohexanoic acid (PFHeA, C5 F11 COOH, 98%) and perfluoropentanoic acid (PFPeA, C4 F9 COOH, 98%) were purchased from Tokyo Kasei Kogyo Co., Ltd. (Tokyo, Japan). Trifluoroacetic acid (TFA, CF3 COOH, 99.0%) was purchased from Aladdin (Shanghai, China). Tridecafluorohexane-1-sulfonic acid potassium salt (PFHS, C6 F13 SO3 K, ≥98.0%), potassium nonafluoro-1-butanesulfonate (PFBS, C4 F9 SO3 K, 98.0%) and heptadecafluorooctanesulfonic acid potassium salt (TFOH, CF3 SO3 K, ≥98.0%) were purchased from Sigma–Aldrich (St. Louis, MO, USA). Analytical-grade reagents of sodium sulfite, ammonia solution, trisodium citrate, sodium nitrite and sodium nitrate were purchased from Sinopharm Chemical Reagent Co., Ltd. (Shanghai,

333

Table 1 The chosen m/z values for PFOA and its possible intermediates quantified by LC/MS in SIM mode. Compound

m/z

Compound

m/z

Compound

m/z

C7 F15 COOH C6 F13 COOH C5 F11 COOH C4 F9 COOH C3 F7 COOH C2 F5 COOH CF3 COOH

413 363 313 263 213 163 113

C7 F14 HCOOH C7 F13 H2 COOH C6 F12 HCOOH C6 F11 H2 COOH C5 F10 HCOOH C5 F9 H2 COOH C4 F8 HCOOH C4 F7 H2 COOH C3 F6 HCOOH C3 F5 H2 COOH C2 F4 HCOOH C2 F3 H2 COOH CF2 HCOOH CFH2 COOH

395 377 345 327 295 277 245 227 195 177 145 127 95 77

C7 F15 SO3 − C6 F13 SO3 − C5 F11 SO3 − C4 F9 SO3 − C3 F7 SO3 − C2 F5 SO3 − CF3 SO3 −

449 399 349 299 249 199 149

China). All the reagents were used as received without further purification. Deionized water was used in the present work. N2 gas with high purity of 99.99% was obtained from Minghui Gas Technology Co., Ltd. (Wuhan, China). 2.2. Degradation experiment The photolytic experiments were conducted in a cylindrical quartz photoreactor with an inner diameter of 58 mm and a length of 158 mm. A 10 W low-pressure mercury lamp (Philips) with emission at 254 nm was used to provide UV irradiation. The reaction temperature was kept at 25 ◦ C by using a cooling water jacket around the reactor. After pH was adjusted with ammonia solution (1:1, V/V) or sulfuric acid (0.1 mol L−1 ), 200 mL of 20.0 ␮mol L−1 PFOA aqueous solution was filled into the reactor. The reaction solution was purged with N2 gas for 30 min to achieve oxygenfree condition. After adding the required amount of sodium sulfite, the solution pH was recorded and regarded as initial pH. Then, the degradation reaction was initiated by switching on the UV lamp. Samples were taken from the reaction solution at predetermined time intervals, followed by analysis. All degradation experiments were conducted in duplicate and the averaged values were presented as the results. 2.3. Analytical methods PFOA and its degradation intermediates were analyzed by Waters 2695 HPLC system coupled with Waters Acquity TQD MS system. The sample separation was performed with a ZORBAX SB-C18 (150 mm × 4.6 mm, 5 ␮m). Mobile phase was composed of methanol and 5 mmol L−1 ammonium acetate (pH 6.0) aqueous solution operated by increasing methanol from 40% to 80% in 10 min, holding on 5 min and then reverting to initial conditions within 2 min. Between two successive samples, the equilibration time was 5 min. The flow rate and injected volume were 0.2 mL min−1 and 10 ␮L, respectively. Column temperature was kept at 40 ◦ C. The electron-spray ionization conditions in the negative ion mode were as follows: capillary voltages 2.5 kV, the cone voltage 12 V, source temperature 120 ◦ C, desolvation temperature 350 ◦ C, and desolvation gas flow 550 L h−1 . Selected ion monitoring (SIM) mode was used for the quantitation of PFOA and its degradation products. The chosen m/z values were shown in Table 1. H2 O2 was measured by using a N,N-diethyl-p-phenylenediamine sulfate (DPD) method [25]. The DPD method is a spectrophotometric method for determination of H2 O2 concentration in the solution, where DPD is oxidized by H2 O2 in the presence of Horseradish peroxidase to form a colored product DPD•+ , which has an absorption maximum at 551 nm wavelength. The concentration

334

Z. Song et al. / Journal of Hazardous Materials 262 (2013) 332–338

Defluorination efficiency / %

100

the photo-conversion of SO3 2− . If the defluorination was conducted under aerated conditions (i.e., in the UV–SO3 2− –air system), the defluorination ratio of PFOA rapidly dropped to about 6.4%, because • oxygen consumes e− aq and H [18]. This is in contrast to the beneficial role of dissolved oxygen in the oxidative degradation of PFOA [10,27–29], and further confirms that the PFOA decomposition in UV–SO3 2− –N2 system proceeded in a reductive manner. − and NO − To illustrate the effect of e− 3 aq , we added NO2 −1 (10 mmol L ) separately into the reaction solution. As shown in Fig. 1b, the addition of NaNO2 and NaNO3 resulted in decreasing the defluorination ratio (in 24 h) from 88.5% to 5.0% and 4.1%, respectively. This is due to the quenching effect of NO2 − and NO3 − on e− aq and H• [24]. It was worth noting that the H• scavenging activity of NO2 − (k = 7.1 × 108 M−1 s−1 ) is about 500 times higher than that of NO3 − (k = 1.4 × 106 M−1 s−1 ). However, at the equal load level, NO2 − and NO3 − caused almost the same inhibition degree on the PFOA defluorination. This indicates that H• made a minor contribution to the PFOA defluorination, possibly because the used alkaline solution (i.e., pH 10.3) is unfavorable to the formation of H• . These results further confirm that the main active species is e− aq in the SO3 2− -mediated photolysis under N2 atmosphere.

(a)

80 2-

UV-SO3 -N2

60

2-

UV-SO3 -Air UV-N2

40

SO32- -N2

20 0

0

6

12

18

24

Time / h

Defluorination efficiency / %

100

(b)

80 60

No scavengers 10mM NaNO3

40

10mM NaNO2

3.2. Effects of main reaction parameters As stated above, e− aq plays a critical role in the PFOA defluorination in the UV–SO3 2− –N2 process. According to the literature survey, the possible main reactions involve in the UV–SO3 2− –N2 process are summarized as follows,

20 0

0

6

12

18

SO3 2− + hv → SO3 •− + eaq − [23]

24

(2)

Time / h Fig. 1. (a) Defluorination of PFOA in different systems. (b) Inhibition effect of e− aq scavengers on the defluorination of PFOA in the UV–SO3 2− –N2 process. Unless otherwise stated, the basic reaction conditions were as follows: initial PFOA concentration, 20.0 ␮mol L−1 ; initial SO3 2− concentration, 10.0 mmol L−1 ; and initial solution pH, 10.3.

of fluoride ions was determined by a fluoride ion selective electrode (Ruosull Co., Ltd., Shanghai). This method provides a detection limit (signal-to-noise ratio of 3) as low as 0.53 ␮mol L−1 of fluoride ions. Defluorination efficiency (DE) was calculated as follows: cF− Defluorination efficiency = × 100% c0 × 15

(1) (␮mol L−1 ),

2SO3 •− + 2SO3 •− + H2 O → S2 O6 2− + SO4 2− + H+ + HSO3 − k3 = 3.1 × 108 M−1 s−1 [30]

SO3 •− + SO3 •− → S2 O6 2−

(3)

k3a /k3 = 0.37 [23]

SO3 •− + SO3 •− + H2 O → SO4 2− + H+ + HSO3 − [23] S2 O6 2− + eaq − → SO3 2− + SO3 •− HSO3 − + eaq − → SO3 2− + H•

k4 = 2 × 105 M−1 s−1 [23]

k5 = 2 × 107 M−1 s−1 [31]

(3a) (3b) (4) (5)

where cF− is the concentration of fluoride ions c0 is the initial concentration of PFOA (␮mol L−1 ), and the factor 15 corresponds to the number of fluorine atoms in one PFOA molecule.

eaq − + H+ → H•

3. Results and discussion

NO2 − + eaq − → (NO2 )•2−

k7 = 6 × 109 M−1 s−1

(7)

3.1. Reductive defluorination of PFOA

(NO2 )•2− + H3 O+  NO2 H•− + H2 O pK a = 3.3

(8)

The decomposition of PFOA was conducted at an initial solution pH 10.3 in the systems of UV–SO3 2− –N2 , UV–N2 , SO3 2− –N2 , and UV–SO3 2− –air, respectively. As shown in Fig. 1a, the direct photolysis for 24 h caused a defluorination ratio as low as 2.3%, due to the weak absorption of PFOA at 254 nm (Fig. S1 in SI). Although SO3 2− itself as a reducing agent can eliminate chlorinated byproducts [26], it alone could not induce any defluorination of PFOA because no fluoride ions were detected after 24 h at dark. However, the combination of UV irradiation and SO3 2− in the deaerated system significantly enhanced the decomposition of PFOA, leading to a defluorination ratio as high as 88.5% within 24 h. This greatly enhanced PFOA defluorination in the UV–SO3 2− –N2 sys• tem is possibly attributed to the generated e− aq and/or H through

(NO2 )•2− + H2 O → NO• + 2OH−

(9)

k6 = 2.3 × 1010 M−1 s−1 [18]

(6)

In the presence of NO2 − [32],

NO• + eaq − → NO−

k10 = 1.1 × 1010 M−1 s−1

(10)

−

In the presence of NO3 [32], NO3 − + eaq − → (NO3 )•2−

k11 = 1 × 1010 M−1 s−1

(NO3 )•2− + H2 O  (NO3 H)•− + OH− (NO3 H)•− → NO2 • + OH− 2NO2 • → N2 O4

pK a = 7.5

k13 = 2.3 × 105 s−1

(11) (12) (13) (14)

Z. Song et al. / Journal of Hazardous Materials 262 (2013) 332–338

100

80 60 40 20 0

80 60 40 20 0

0

6

12

18

24

0

6

27

(c)

12

18

24

Time / h

Time / h 0.35

(b)

pH 6.0 pH 8.1 pH 9.3 pH 10.3

(a)

3.0 mM 1.0 mM 0.5 mM

0.20

21

(d)

18

0.20 0.15 9 RQSC Defluorination efficiency

0.10 0.05 0.00 0

5

10

15

20

2-

-1

RQSC Defluorination efficiency

0.15

14

0.10 7 0.05

0.00

0

6

-1

Concentration of SO3 / mmol L

RQSC / µ mol L

0.25

Defluorination efficiency / %

RQSC

/ µ mol L

-1

0.30

7

8

9

10

Defluorination efficiency / %

20.0 mM 10.0 mM 5.0 mM

Defluorination efficiency / %

Defluorination efficiency / %

100

335

0 11

pH

Fig. 2. Effects of (a) SO3 2− concentration and (b) initial solution pH on PFOA defluorination in the UV–SO3 2− –N2 system. Dependences of RQSC and defluorination efficiency (within 0.5 h) on the initial SO3 2− concentration (c) and solution pH (d). Here, RQSC is the relative quasi-stationary concentration of e− aq .

N2 O4 + H2 O → NO2 − + NO3 − + 2H+ SO2 ·H2 O  HSO3 − + H+ HSO3 −  SO3 2− + H+

(15)

pK a1 = 1.76 [33]

(16)

pK a2 = 7.20 [33]

From Eq. (2), we have the rate of by Eq. (18) [34],





r = ˚I0 bεSO2− cSO2− 1 − 10−A /AV 3

k15 = 1 × 103 s−1

e− aq

(17)

generation (r) as indicated

(M s−1 )

(18)

3

where ˚ (mol Einstein−1 ) is the quantum efficiency for SO3 2− pho−1 tolysis or e− aq formation, I0 (Einstein s ) is the photon flux entering the solution from the UV source, b (cm) is the effective path length of light, εSO2− ((mol L−1 )−1 cm−1 ) represents the molar extinction 3

coefficient of SO3 2− , cSO2− (mol L−1 ) is the concentration of SO3 2− , 3

A is the absorbance of solution, and V (L) is solution volume. By considering the e− aq -consuming reactions listed above, the relative 2− –N quasi-stationary concentration of e− 2 aq (RQSC) in the UV–SO3 system in the absence of PFOA may be calculated with Eq. (19) (more details for the derivation process can be found in Text S1, SI), RQSC =

AV



˚I0 bεSO2− (1 − 10−A )Ka1 Ka2 csulfite 3



k5 Ka1 csulfite 10−pH + k6 10−pH (10−2pH + Ka1 10−pH + Ka1 Ka2 )

(19)

will result in a higher reductive degradation rate of both PFOA and its degradation intermediates. The effects of SO3 2− concentration and solution pH on the PFOA defluorination were investigated experimentally. As shown in Fig. 2a, the defluorination ratio of PFOA within 6 h was increased from 5.6% to 68.6% when SO3 2− concentration was increased from 0.5 to 20.0 mmol L−1 . This is attributed to that the increased SO3 2− concentration causes a roughly linearly increasing in the generation rate of e− aq according to Eq. (20). Fig. 2b shows that the defluorination of PFOA was promoted obviously with increasing pH from 6.0 to 10.3. This indicates that the e− aq -mediated reduction of PFOA is more efficient under alkaline conditions, being in agreement with previous reports [20,24]. Fig. 2c and d compared the initial defluorination efficiency of PFOA (within 0.5 h) with the corresponding 2− –N as a function RQSC value of e− 2 aq in the system of UV–SO3 2− of either SO3 concentration or solution pH. It was clearly seen that both the PFOA defluorination and the e− aq RQSC were increased with increasing SO3 2− concentration or solution pH. Moreover, as SO3 2− concentration and solution pH changed, the PFOA defluorination showed a roughly similar trends with the calculated values − of e− aq RQSC. This further demonstrates that eaq is the main reactive species that contribute to the PFOA defluorination. Here, it should be noted that the PFOA defluorination by the UV–SO3 2− –N2 process should be conducted at about pH 10.

At pH > 9, Eq. (19) is simplified to Eq. (20), RQSC =

˚I0 bεSO2− cSO2− (1 − 10−A ) 3

3

AVk6 10

−pH

3.3. Degradation intermediates (20)

2− It is certain that the RQSC of e− aq is mainly affected by initial SO3 concentration and solution pH. Apparently, a higher e− RQSC value aq

Fig. 3a compared the degradation removal and defluorination ratio of PFOA. At a given reaction time, the defluorination ratio was lower than the degradation removal, hinting that some

336

Z. Song et al. / Journal of Hazardous Materials 262 (2013) 332–338

Besides PFCA intermediates, other two groups of F-containing compounds were also detected by using LC/MS (Fig. S5 in SI). One group was the less fluorinated carboxylic acids, including C7 F14 HCOOH, C7 F13 H2 COOH, C6 F12 HCOOH, C5 F10 HCOOH, C4 F8 HCOOH and CF2 HCOOH. The generation of this group of intermediates clearly demonstrates the mechanism of the reductive defluorination of PFOA, in which the C F bond adjacent to the carboxylic group of a PFOA molecule is cleaved by hydrated electrons (see Section 3.4 for more details). This is different from that PFOA is degraded in a stepwise manner with the removal of CF2 unit from its molecule during oxidative degradation. The other group was the fluorinated alkyl sulfonates, including C7 F15 SO3 − , C6 F13 SO3 − , C5 F11 SO3 − , C4 F9 SO3 − , and C3 F7 SO3 − . Like the PFCA intermediates, most of the latter two groups of intermediates were first accumulated and then degraded with increasing the reaction time. This implies that the degradation rates of these compounds were slower than other groups of intermediates (also see Text S4 in SI). However, it should be noted that the concentrations of all three groups of F-containing degradation intermediates were lower than 0.02 ␮mol L−1 , being much smaller than the initial concentration of parent PFOA (20.0 ␮mol L−1 ). This result is in good agreement with the defluorination efficiency as high as 88.5% at 24 h.

(a)

Percentage / %

100 80 60 Degradation efficiency Defluorination efficiency

40 20 0

0

6

12

18

24

Time / h C6F13COOH

0.20

C4F9COOH

0.15

0.04

CFCOOH

Concentration / µ mol L

-1

Concentration / µ mol L

(b)

C5F11COOH

CFCOOH

0.03

CFCOOH

0.02

0.10

3.4. Decomposition mechanism

0.01

0.00 0

0.05

0.00

0

6

6

12

12

Time / h

18

18

As discussed above, the main active species responding for PFOA defluorination are the photo-generated hydrated electrons in the UV–SO3 2− –N2 process. Due to their strong electron affinity, the hydrated electrons initially access and attack the fluorine atom at ␣position, leading to the degradation of PFOA in a stepwise manner (Fig. 4). This accounts for the observation that the less fluorinated carboxylic acids C7 F14 HCOOH and C7 F13 H2 COOH were detected at the initial reaction period, and the concentration of C7 F14 HCOOH was higher than that of C7 F13 H2 COOH (Fig. S5a). The generated C7 F13 H2 COOH is further transformed to C6 F13 COOH through reactions (21) and (22),

24

24

Time / h Fig. 3. (a) Irradiation-time dependences of degradation removal and defluorination ratio of PFOA. (b) Concentrations of shorter chain PFCA intermediates generated during the reductive degradation of PFOA.

F-containing intermediates were generated during the PFOA defluorination. Indeed, shorter chain perfluorocarboxylic acids (PFCAs) with 2–7 carbon atoms were identified by LC/MS. As shown in Fig. 3b, the concentrations of these intermediates firstly increased, and then decreased, indicating that they were generated and then further decomposed. Besides, the maximum concentration of the intermediates occurred at about 0.5 h for C6 F13 COOH, C5 F11 COOH and C4 F9 COOH, and at about 1.5 h for C3 F7 COOH, C2 F5 COOH and CF3 COOH, respectively. These implied that both PFOA and its intermediates were decomposed in a stepwise manner. F C6F13

C

COOH

F

C7 F13 H2 COOH → • C6 F13 + : CH2 + • COOH

(21)

•C

(22)

6 F13

The concentration of C6 F12 HCOOH was higher than that of C6 F11 H2 COOH (Fig. S5b), thus, it can be concluded that like the removal pathway of CF2 unit from its parent C7 F15 COOH, the generated C6 F13 COOH is further decomposed to C6 F12 HCOOH, then to C6 F11 H2 COOH, and so forth to C5 F11 COOH. This defluorination pathway is very different from that in the system of UV–N2 . During the direct photolysis, the C C bond between C7 F15 − and COOH of H

+ eaq-

C6F13

- F-

+ • COOH → C6 F13 COOH

C

COOH

H

+ eaq-

C6F13

- F-

F

COOH

H

m/z 395

m/z 413

C

m/z 377

- CH2 hv Defluorination

F C5F11

C

COOH

F m/z 363

F-

+

Defluorination products

Fig. 4. Possible pathway of PFOA degradation in the UV–SO3 2− –N2 system. Due to its strong electronegativity, fluorine atoms act as reaction centers during the reductive degradation. Hence, hydrated electrons (e− aq ) as the main active species initially access and attack the fluorine atom at ␣-position, leading to the degradation of PFOA through less fluorinated carboxylic acids in a stepwise manner.

Z. Song et al. / Journal of Hazardous Materials 262 (2013) 332–338

PFOA was destructed firstly, initiating the following reactions (Eqs. (23)–(26)) [9,10], C7 F15 COOH + hv → C7 F15 • + • COOH

(23)

C7 F15 • + H2 O → C7 F15 OH + H•

(24)

C7 F15 OH → C6 F13 COF + H+ + F−

(25)

C6 F13 COF + H2 O → C6 F13 COOH + H+ + F−

(26)

21177044), the National High Technology Research and Development Program of China (863) Program (Grant No. 2012AA06A304), the Scientific Research Key Project of Hubei Provincial Department of Education (Project No. D20121503), and the Fundamental Research Founds for the Central University of China (Grant No. CZZ11008). The authors also acknowledge the very helpful discussion with the reviewers. Appendix A. Supplementary data

The formed shorter chain intermediate C6 F13 COOH was further decomposed in a same manner by stepwise losing a CF2 unit. However, the generation of the PFCAs bearing shorter chains through the direct photolysis is ignorable in comparison with that through the e− aq reduction route in the sulfite/UV process. Moreover, the sulfonated intermediates might be formed from the reaction of Cn F2n+1 • radicals with SO3 •− (Eq. (27)), which was generated from the UV excitation of sulfite solution: Cn F2n+1 • + SO3 •− → Cn F2n+1 SO3 −

337

(27)

e− aq

In the reduction route, C6 F13 • radicals are the primary Cn F2n+1 • species during the defluorination of PFOA (Eq. (21)) but without the formation of C7 F15 • , while C7 F15 • radicals are firstly formed through the direct photolysis. It was found that the generation and accumulation of C7 F15 SO3 − was much less than C6 F13 SO3 − in the UV–SO3 2− –N2 system (Fig. S5c). This again confirmed that the defluorination of PFOA mainly induced by the e− aq reduction route (Fig. 1a) in the UV–SO3 2− –N2 system. As for the further degradation of the fluorinated sulfonate intermediates, we may refer to the reports on the degradation of perfluorinated alkyl sulfonates in literature. Yamamoto et al. [35] reported that the photodegradation of PFOS was started from the cleavage of C S bonds between C8 F17 − and SO3 − with a rate constant of 0.93 days−1 . Park et al. [21] indicated that the e− aq firstly attacked the fluorine atoms across the fluorocarbon tail but not being limited to the ␣-position. In the present work, we did not intend to investigate the degradation mechanism of the fluorinated sulfonate intermediates. By briefly making that all the intermediates will be completely mineralized to CO2 , SO4 2− and F− after the long time reaction, we summarized the possible pathways for the photo-reductive degradation of PFOA in the UV–SO3 2− –N2 system as shown in Fig. 4. 4. Conclusions Efficient photo-reductive defluorination of PFOA was achieved by using SO3 2− as a mediator. The photolysis of SO3 2− under alkaline conditions produced hydrated electrons and SO3 •− . The former as the main active species is responsible for the PFOA defluorination, while the later participates in the degradation process, resulting in the formation of some perfluorinated alkyl sulfonates as intermediates. The enhancement of the defluorination of PFOA with increasing SO3 2− concentration and pH was derived from the production of more hydrated electrons, as confirmed by theoretical reaction kinetic analysis. A mechanism was proposed for the defluorination of PFOA in the investigated system. This study provides a new reductive method to efficiently decompose PFOA and other PFCAs under mild conditions, i.e., room temperature and atmospheric pressure. Acknowledgements The authors acknowledge the funding support from the National Natural Science Foundation (Grants Nos. 21077037, 21107027 and

Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.jhazmat.2013. 08.059. References [1] G.W. Olsen, J.M. Burris, J.H. Mandel, L.R. Zobel, Serum perfluorooctane sulfonate and hepatic and lipid clinical chemistry tests in fluorochemical production employees, J. Occup. Environ. Med. 41 (1999) 799–806. [2] T. Wang, Y.W. Wang, C.Y. Liao, Y.Q. Cai, G.B. Jiang, Perspectives on the inclusion of perfluorooctane sulfonate into the Stockholm convention on persistent organic pollutants, Environ. Sci. Technol. 43 (2009) 5171–5175. [3] U.S. Environmental Protection Agency Office of Pollution Prevention and Toxics Risk Assessment Division, Preliminary risk assessment of the developmental toxicity associated with exposure to perfluorooctanoic acid and its salts, 2003. [4] B.D. Key, R.D. Howell, C.S. Criddle, Fluorinated organics in the biosphere, Environ. Sci. Technol. 31 (1997) 2445–2454. [5] J.P. Giesy, K. Kannan, Perfluorochemical surfactants in the environment, Environ. Sci. Technol. 36 (2002) 146A–152A. [6] J.P. Giesy, K. Kannan, Global distribution of perfluorooctane sulfonate in wildlife, Environ. Sci. Technol. 35 (2001) 1339–1342. [7] M. Houde, J.W. Martin, R.J. Letcher, K.R. Solomon, D.C.G. Muir, Biological monitoring of polyfluoroalkyl substances: a review, Environ. Sci. Technol. 40 (2006) 3463–3473. [8] R.R. Giri, H. Ozaki, T. Okada, S. Taniguchi, R. Takanami, Factors influencing UV photodecomposition of perfluorooctanoic acid in water, Chem. Eng. J. 180 (2012) 197–203. [9] H. Hori, E. Hayakawa, H. Einaga, S. Kutsuna, K. Koike, T. Ibusuki, H. Kiatagawa, R. Arakawa, Decomposition of environmentally persistent perfluorooctanoic acid in water by photochemical approaches, Environ. Sci. Technol. 38 (2004) 6118–6124. [10] H. Hori, A. Yamamoto, E. Hayakawa, S. Taniyasu, N. Yamashita, S. Kutsuna, Efficient decomposition of environmentally persistent perfluorocarboxylic acids by use of persulfate as a photochemical oxidant, Environ. Sci. Technol. 39 (2005) 2383–2388. [11] B.B. Wang, M.H. Cao, Z.J. Tan, L.L. Wang, S.H. Yuan, J. Chen, Photochemical decomposition of perfluorodecanoic acid in aqueous solution with VUV light irradiation, J. Hazard. Mater. 181 (2010) 187–192. [12] H.Q. Tang, Q.Q. Xiang, M. Lei, J.C. Yan, L.H. Zhu, J. Zou, Efficient degradation of perfluorooctanoic acid by UV-Fenton process, Chem. Eng. J. 184 (2012) 156–162. [13] S.C. Panchangam, A.Y.C. Lin, K.L. Shaik, C.F. Lin, Decomposition of perfluorocarboxylic acids (PFCAs) by heterogeneous photocatalysis in acidic aqueous medium, Chemosphere 77 (2009) 242–248. [14] Y.C. Chen, S.L. Lo, J. Kuo, Effects of titanate nanotubes synthesized by a microwave hydrothermal method on photocatalytic decomposition of perfluorooctanoic acid, Water Res. 45 (2011) 4131–4140. [15] C. Song, P. Chen, C.Y. Wang, L.Y. Zhu, Photodegradation of perfluorooctanoic acid by synthesized TiO2 –MWCNT composites under 365 nm UV irradiation, Chemosphere 86 (2012) 853–859. [16] X.Y. Li, P.Y. Zhang, L. Jin, T. Shao, Z.M. Li, J.J. Cao, Efficient photocatalytic decomposition of perfluorooctanoic acid by indium oxide and its mechanism, Environ. Sci. Technol. 46 (2012) 5528–5534. [17] Z.M. Li, P.Y. Zhang, T. Shao, X.Y. Li, In2 O3 nanoporous nanosphere: a highly efficient photocatalyst for decomposition of perfluorooctanoic acid, Appl. Catal. B: Environ. 125 (2012) 350–357. [18] G.V. Buxton, C.L. Greenstock, W.P. Helman, A.B. Ross, Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals (• OH/• O− ) in aqueous solution, Phys. Chem. Ref. Data 17 (1988) 513–886. [19] L. Huang, W.B. Dong, H.Q. Hou, Investigation of the reactivity of hydrated electron toward perfluorinated carboxylates by laser flash photolysis, Chem. Phys. Lett. 436 (2007) 124–128. [20] Y. Qu, C.J. Zhang, F. Li, J. Chen, Q. Zhou, Photo-reductive defluorination of perfluorooctanoic acid in water, Water Res. 44 (2010) 2939–2947. [21] H. Park, C.D. Vecitis, J. Cheng, W. Choi, B.T. Mader, M.R. Hoffmann, Reductive defluorination of aqueous perfluorinated alkyl surfactants: effects of ionic headgroup and chain length, J. Phys. Chem. A 113 (2009) 690–696. [22] E.D. Williams, I. Doniach, O. Bjarnason, W. Michie, Thyroid cancer in an iodide rich area. A histopathological study, Cancer 39 (1977) 215–222.

338

Z. Song et al. / Journal of Hazardous Materials 262 (2013) 332–338

[23] M. Fischer, P. Warneck, Photodecomposition and photooxidation of hydrogen sulfite in aqueous solution, J. Phys. Chem. 100 (1996) 15111–15117. [24] X.C. Li, J. Ma, G.F. Liu, J.Y. Fang, S.Y. Yue, Y.H. Guan, L.W. Chen, X.W. Liu, Efficient reductive dechlorination of monochloroacetic acid by sulfite/UV process, Environ. Sci. Technol. 46 (2012) 7342–7349. [25] H. Bader, V. Sturzenegger, J. Hoigné, Photometric method for the determination of low concentrations of hydrogen peroxide by the peroxidase catalyzed oxidation of N,N-diethyl-p-phenylenediamine (DPD), Water Res. 22 (1988) 1109–1115. [26] J.P. Croue, D.A. Reckhow, Destruction of chlorination byproducts with sulfite, Environ. Sci. Technol. 23 (1989) 1412–1419. [27] Y. Wang, P.Y. Zhang, G. Pan, H. Chen, Ferric ion mediated photochemical decomposition of perfluorooctanoic acid (PFOA) by 254 nm UV light, J. Hazard. Mater. 160 (2008) 181–186. [28] H. Hori, Y. Takano, K. Koike, K. Takeuchi, H. Einaga, Decomposition of environmentally persistent trifluoroacetic acid to fluoride ions by a homogeneous photocatalyst in water, Environ. Sci. Technol. 37 (2003) 418–422.

[29] R. Dillert, D. Bahnemann, H. Hidaka, Light-induced degradation of perfluorocarboxylic acids in the presence of titanium dioxide, Chemosphere 67 (2007) 785–792. [30] P. Warneck (Ed.), Transport and Chemical Transformation of Pollutants in the Troposphere, vol. 2, Heterogeneous and Liquid Phase Processes, SpringerVerlag, Berlin, 1996. [31] E. Hayon, A. Treinin, J. Wilf, Electronic spectra, photochemistry, and autoxidation mechanism of the sulfite-bisulfite-pyrosulfite systems. The SO2 − , SO3 − , SO4 − , and SO5 − radicals, J. Am. Chem. Soc. 94 (1972) 47–57. [32] M.G. Gonzalez, E. Oliveros, M. Wˇorner, A.M. Braun, Vacuum-ultraviolet photolysis of aqueous reaction systems, J. Photochem. Photobiol. C 5 (2004) 225–246. [33] H.V. Tartar, H.H. Garretson, The thermodynamic ionization constants of sulfurous acid at 25◦ , J. Am. Chem. Soc. 63 (1941) 808–816. [34] J.C. Crittenden, S.M. Hu, D.W. Hand, S.A. Green, A kinetic model for H2 O2 /UV process in a completely mixed batch reactor, Water Res. 33 (1999) 2315–2328. [35] T. Yamamoto, Y. Noma, S. Sakai, Y. Shibata, Photodegradation of perfluorooctane sulfonate by UV irradiation in water and alkaline 2-propanol, Environ. Sci. Technol. 41 (2007) 5660–5665.