Degradation performance and mechanism of decabromodiphenyl ether (BDE209) by ferrous-activated persulfate in spiked soil

Accepted Manuscript Degradation performance and mechanism of decabromodiphenyl ether (BDE209) by ferrous-activated persulfate in spiked soil Hongjiang Peng, Wei Zhang, Lin Liu, Kuangfei Lin PII: DOI: Reference:

S1385-8947(16)31221-9 http://dx.doi.org/10.1016/j.cej.2016.08.129 CEJ 15695

To appear in:

Chemical Engineering Journal

Received Date: Revised Date: Accepted Date:

5 May 2016 27 August 2016 29 August 2016

Please cite this article as: H. Peng, W. Zhang, L. Liu, K. Lin, Degradation performance and mechanism of decabromodiphenyl ether (BDE209) by ferrous-activated persulfate in spiked soil, Chemical Engineering Journal (2016), doi: http://dx.doi.org/10.1016/j.cej.2016.08.129

This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

1

Degradation performance and mechanism of decabromodiphenyl

2

ether (BDE209) by ferrous-activated persulfate in spiked soil

3 4

Hongjiang Penga,b , Wei Zhanga,b,*, Lin Liua,b, Kuangfei Lina,b

5 6

a

7

on Chemical Process, East China University of Science and Technology, Shanghai 200237, China

8

b

9

Technology, Shanghai 200237, China

State Environmental Protection Key Laboratory of Environmental Risk Assessment and Control

School of Resource and Environmental Engineering, East China University of Science and

10 11

AB STRAC T

12

This study first investigated the degradation performance of BDE209 using ferrous activated

13

persulfate-based advanced oxidation process. The results indicated that a lower pH would result in

14

a greater increase in the BDE209 removal efficiency, and the maximum removal efficiency was

15

obtained at pH=3.0. The effects of sodium persulfate (PS) dosage and molar ratio of PS/Fe(II)

16

were also determined, and 0.2 mol L-1 and 2:1 were the best conditions, where the removal of

17

BDE209 in soil could reach to about 53% after 6 h. Additionally, hydroxylamine (HA) was firstly

18

introduced to a PS/Fe(II) system, and resulted in a large enhancement of BDE209 removal

∗ Corresponding author at: School of Resource and Environmental Engineering, East China University of Science and Technology, Shanghai 200237, China. Tel.: +86 21 64253244. Fax: +86 21 64253988. E-mail address: [email protected] (W. Zhang). 1

19

efficiency. Compared to the controls, the degradation rate increased by 13% with the ratio of

20

HA/Fe(II) 4/1, which might be because HA accelerated the transformation from Fe(III) to Fe(II).

21

Additionally, 9 intermediate products during iron-activated persulfate oxidation process were

22

identified, and a possible reaction mechanism was further proposed.

23

Keywords: BDE209; sodium persulfate; intermediate product; sulfate radical; hydroxylamine.

24 25

1. Introduction

26

Due to excellent flame retardant properties, polybrominated diphenyl ethers (PBDEs) have

27

been used as flame retardants in a variety of commercial products[1]. PBDEs can enter into the

28

environment easily during their production, assumption and treatment processes, because they are

29

not chemically bonded but just added into the products. PBDEs may reach soil via wet deposition

30

and dry deposition during their long-range atmospheric transport, and tend to be absorbed in soil

31

strongly due to their abilities of persistence and lipophilicity[2]. China has become the largest

32

dumping site of e-wastes in the world. It is estimated that the global 50~80% of e-waste imports to

33

Asia via different routes, as well as 50~70% of them is destined for China[3]. BDE209 dominates

34

PBDEs homologues, and its consumption in China ranges from about 20000-40000 tons per year,

35

while until now there is no restriction to BDE209 use[4]. Luo et al.[5] reported that BDE209

36

contents in farmland soils from an e-waste recycling workshop were 0.0691~6.319 µg g-1.

37

For decades, many efforts to degrade PBDEs have been taken, such as TiO2 photocatalysis[6],

38

chemical reductions with zero-valent irons[7] and nanoscale zero-valent irons[8]. In recent years,

39

persulfate anion (S2O8 2-) was drawing increasing attentions due to its strong oxidation-reduction

40

potential (E0∼2.01 V) and that can be converted to even stronger sulfate radical with higher 2

41

oxidation-reduction potential (E0∼2.60 V)[9, 10]. The SO4−• can be effectively generated by

42

different methods to activated the S2O82-, such as heat[11, 12], UV light[13-15], alkaline pH[16],

43

and transition metals (Men+)[17]. In addition, the S2O82- can persist in underground systems longer

44

than other oxidants, such as ozone and hydrogen peroxide[18]. For in situ applications, ferrous and

45

ferric ions are the most widely used metal activators due to their natural abundance[19]. Compared

46

to thermal activation (33.5 kcal mol-1), the activation by Fe(II) requires a relatively lower energy

47

(i.e., 14.8 kcal mol-1) , which poses a great potential to destruct the contaminants by rapidly

48

generating sulfate radical[20]. Persulfate anion activated by Fe(II) to generate SO4−• and sulfate, as

49

depicted by Eq. (1)[21].

50

S2O82- + Fe2+ → Fe3+ + SO42- + SO4−•

k=15.33 M-1 s-1

51

However, Fe(II) activated PS technology still has some intrinsic drawbacks. For example, the

52

fast conversion of Fe(II) into Fe(III) will result in a rapid decline of PS activation efficiency; In

53

addition, excessive amount of Fe(II) can also act as an effective scavenger of SO4−• at its high

54

concentration as expressed by Eq. (2), which leads to the decline of oxidation efficiency[10]. k=4.9×109 M-1 s-1

(1)

55

SO4−• + Fe2+ → Fe3+ + SO42-

56

The main defect in PS/Fe(II) system is slow conversion rate from Fe(III) to Fe(II). In order to

57

overcome the drawback, we considered some reducing agents with low reaction rates to reactive

58

species. Hydroxylamine (NH2OH, HA), a well-known reducing agent with a strong reduction

59

ability to transfer Fe(III) to Fe(II), has been employed in many applications such as total iron

60

determination[22]. In addition, the end-products of NH2OH are inorganic substances, such as N2,

61

N2O, NO2-, and NO3-[23-25]. Due to its strong reducing ability to transfer Fe(III) to Fe(II) and low

3

(2)

62

rate constants with SO4-•[26] and OH•[27], HA was introduced into a PS/Fe(II) system to facilitate

63

the degradation efficiency of BDE209 in this study.

64

The present research aimed to: (1) evaluate the technical feasibility of BDE209 removal by

65

ferrous-activated persulfate in soil; (2) explore the impacts of oxidant dosage, initial soil pH, initial

66

hydroxylamine dosage and BDE209 concentration on the removal of BDE209; (3) analyze the

67

intermediates of BDE209 during the oxidative process, and then propose possible degradation

68

pathways.

69 70

2. Materials and methods

71

2.1. Chemicals and soils

72

BDE209 (purity > 98.0%) was obtained from J&K Scientific Ltd., Shanghai, China.

73

Hydroxylamine sulfate (HA, > 99.0%) and sodium persulfate (PS, > 99.5%) were purchased from

74

Sinopharm Chemical Reagent Co., Ltd., Shanghai, China. Toluene, dichloromethane and n-hexane

75

were obtained from Lingfeng Chemical Co., Ltd., Shanghai, China. All organic solvents used in

76

the experiments were analytical grade. All the solutions were freshly prepared with deionized

77

water before each run.

78

Uncontaminated soil was collected from ECUST (East China University of Science and

79

Technology), Shanghai, China. Soil samples dried naturally at room temperature, then sieved with a

80

2-mm mesh to remove stones and debris, and then stored in plastic bags for further experiments.

81

The pH of the soil was 6.5 as determined by USEPA Method 9045D. The soil sample was

82

characterized as a silty clay loam, with 6.5% of organic matter content.

4

83

The soil was spiked with 1:9 (v/v) toluene/n-hexane solutions that contained a known amount

84

of BDE209, and then the spiked soil was placed in a fume hood to eliminate the solvents. BDE209

85

concentration in the prepared soil was 10 mg kg-1.

86 87

2.2. Experimental procedures

88

The degradation experiments were conducted in 50-mL tubes. Before the oxidation process, all

89

vessels were rinsed with n-hexane and then with deionized water. Prior to each experiment, 5.0 g

90

spiked soil was added to the tubes, followed by the addition of 2.5 mL PS and 2.5 mL ferrous sulfate

91

solutions, respectively. After the addition of the two chemicals, the tubes were kept in a water bath

92

oscillator maintained at 25 oC in the dark and was shaken at 150 rpm for reaction with different

93

periods (0, 0.5, 1, 1.5, 2, 4, and 6 h). At each sampling point, all the soils in the tubes were sampled.

94

At pre-specified time intervals, the sample vials were immediately put into an ice-water bath for 10

95

minutes to quench the reaction then put into a -70 o C refrigerator for 2 hours, and then transferred in

96

a freeze-dryer for further 48 hours. All treatments were conducted in triplicates.

97 98

2.3. Samples pretreatment

99

After freeze dried for 48 h, the soil samples were then ultrasonic assisted extraction for 2 times

100

with 50 mL high purity dichloromethane/n-hexane (2:3, v/v) for 30 min. The extracts were

101

concentrated to approximately 4~5 mL by a rotary evaporator (Buchi R-210, Switzerland), and then

102

passed through a PTFE membrane. The concentrated extracts were evaporated to 2 mL under a

103

gentle stream of nitrogen gas.

104 5

105

2.4. Analytical methods

106

BDE209 was determined using Agilent GC/ECD (7890A) equipped with a capillary column

107

(J&W, Scientific, 30.0 m×0.32 mm×0.25 µm, HP-5). The oven temperature was programmed

108

starting at 110 oC for 1 min, then increased to 260 o C at a rate of 40 oC min-1 and then increased to

109

320 oC at a rate of 10 oC where it was hold for 10 min. The temperature of the injector and detector

110

were 280 oC, respectively. High purity nitrogen was used as the carrier gas and the ECD makeup gas

111

with a constant flow of 3.0 mL min-1 and 20 mL min-1, respectively.

112

Agilent GC/MS (7890A/5975C) equipped with a capillary column (J&W, Scientific, 15

113

m×0.25 mm×0.1 µm, DB-5HT) was used to determine potential BDE209 degradation

114

intermediates. The MSD mass spectrometer equipped with a negative chemical ionization (NCI)

115

source or an electron impact (EI). When MS used a NCI mode, both the quadrupole and source

116

temperatures were maintained at 150 oC, while MS used an EI mode, the quadrupole and source

117

temperatures were kept at 150 and 230 oC, respectively. The injection was in splitless mode with

118

exactly 1 µL solution, and the high purity Helium (99.999%) was applied as the carrier gas at a

119

constant flow of 1.0 mL min-1. The oven temperature was programmed starting at 110 oC for 1 min,

120

then increased to 320 oC at a rate of 8 o C min-1 and held for 8 min. The MS interface temperature

121

was 300 oC. The recovery rates ranged from 83.1% to 107.3%, and it should be noted that all values

122

were not corrected by this coefficient.

123 124

2.5. Data analysis

6

125

Each treatment was conducted in triplicates and the results of the analysis were presented as the

126

mean with a standard deviation less than 5%. The figures were generated using origin 8.0

127

(OriginLab, Northampton, MA, USA).

128 129

3. Results and discussion

130

3.1. BDE209 oxidation at different ferrous ion concentrations

131

As shown in Fig. 1, different molar ratios of Fe(II)/PS (0.1/1, 0.2/1, 0.5/1, 1/1 and 2/1,

132

respectively) were applied. The results show that when the Fe(II)-to-PS molar ratios increased

133

from 0.1/1 to 0.5/1, the removal efficiency of BDE209 were 29%, 33% and 53%, with the reaction

134

rate constants of 0.11, 0.16 and 0.48 (h-1), respectively, indicating that the removal rates of BDE209

135

could be significantly enhanced by increasing ferrous ion concentrations within the reasonable

136

Fe(II)-to-PS molar ratios. However, the degradation efficiency of BDE209 began to decline as the

137

amount of Fe(II) increased continuously. Additionally, when the molar ratio of Fe(II)-to-PS over

138

0.5/1 showed no enhancement in BDE209 destruction: The removal rates dropped from 53% to

139

45% and 40%, and the reaction rate constants declined from 0.48 to 0.26 and 0.21 (h-1), accordingly.

140

The reason for this trend should be that the ferrous ion acted as a scavenger for sulfate free

141

radicals (As shown in reaction (2)). In addition, the ferrous ion could also decompose persulfate

142

anion according to reaction (3). Thus, excess amount of ferrous ion would lead to the declination

143

of BDE209 removal efficiency. Similar trend was also observed by Chen et al.[28], who reported

144

that the removal rates of MTBE decreased dramatically with excessive ferrous ions due to the

145

competition for sulfate free radicals between ferrous ions and MTBE. Additionally, as shown in

146

Eq. (3), the stoichiometric ratio between PS and Fe(II) was 2. Higher initial Fe(II) contents meant 7

147

that the ratios of ∆Fe(II)/∆PS were closer to 2. Smaller molar ratios meant less consumption or

148

scavenging of Fe(II) by sulfate radicals (see Eq. (2)); as a consequence, sulfate radicals were

149

mostly used for BDE209 destruction, and higher removal efficiency would be observed. In this

150

study, the results suggested that the optimal molar ratio of Fe(II)/PS was 0.5/1.

151

S2O82- + 2Fe2+ → 2Fe3+ + 2SO42-

(3)

152 153

3.2. BDE209 oxidation at different persulfate concentrations

154

The effects of initial persulfate dosage on the BDE209 degradation were investigated within

155

the ranges of 0.01-1.0 M, fixed the Fe(II)/PS molar ratio of 0.5/1 at 25 o C. Fig. 2 shows that

156

BDE209 removal rate increased a lot as initial PS concentrations increased from 0.01 M to 0.2 M,

157

and the removal of BDE209 after 360 min were from 28% to 53%. Unfortunately, when the PS

158

concentrations increased continuously from 0.2 to 1.0 M, only a little enhancement was observed,

159

and the removal of BDE209 ranged from 53% to 55%, respectively. It should be noted that, at a low

160

level of PS condition, increasing the concentration of oxidant is essential for generating a higher

161

level of sulfate radical to overcome the competition of other organic materials and inorganic

162

ions[19]. However, at higher persulfate concentrations, more SO4−• would be generated,

163

correspondingly. Excessive SO4−• might work as scavengers for S2O82− or SO4−• according to Eqs.

164

(4) and (5), respectively[29, 30].

165

SO4−• + S2O82− → SO42− + S2O8−•

166

SO4−• + SO4−• → S2O82−

8

k=6.1×105 M-1 s-1

(4)

k=4.0×108 M-1 s-1

(5)

167

Over all, considering the effects of BDE209 oxidation and the dosage of oxidant, the optimal

168

PS concentration was finally considered as 0.2 M as the initial persulfate contents range of 0-1.0 M

169

in the present study.

170 171

3.3. BDE209 oxidation at different initial pH values

172

The investigation of the effects of varying initial pH conditions on the degradation of

173

BDE209 by the Fe(II)/PS process was carried out for pH 3.0, 5.0, 7.0 and 9.0. The soil pH was

174

adjusted by using 1 M sulfuric acid or 1 M sodium hydroxide. No buffer solution was used in our

175

experiment in order to avoid it to react with free radicals. And the pH barely kept constant over the

176

course of the experiment.

177

Fig. 3 shows the removal profile of BDE209 by the Fe(II)/PS process, and the pseudo

178

first-order kinetic rate constants (kobs) at different pH levels were determined as presented in Table

179

1. The degradation reaction of BDE209 was obviously dependent on the pH. The degradation rate

180

constants of BDE209 decreased from 0.48 to 0.34 h-1 as the slurry pH increased from 3.0 to 9.0.

181

The rate constants of BDE209 degradation in the acidic conditions were higher than in the alkaline

182

regimes, and the maximum degradation rate occured at pH=3.0. The reason may be that, under

183

acidic conditions, the efficiency of sodium sulfate conversion to sulfate radical will increase due to

184

the acid catalysis[31]. As a result, at lower pH values, higher efficiency of BDE209 degradation

185

can be observed.

186 187

3.4. BDE209 oxidation at different initial HA dosages

9

188

The main disadvantage of the Fe(II)/PS system was the accumulation of Fe(III), which may

189

further reduce degradation efficiency of the target contaminants. In order to overcome this

190

drawback, some reducing agents can be added to improve the removal rate. Hence, HA was

191

applied to aim at accelerating the recovering of Fe(II) (as shown in Eq. (6))[32]. Moreover, in order

192

to avoid introducing Cl−, we used hydroxylamine sulfate as the source of NH3OH+. Thus, to obtain

193

further insights into the role of HA, different dosages of HA were explored.

194

NH3 OH+ + Fe3+ → Fe2+ + nitrogenous products

195

As shown in Fig. 4, without adding hydroxylamine, when the reaction was conducted for 2

196

hours, no further increase in the rate of BDE209 degradation due to the Fe(III) accumulation and

197

the ultimate removal rate was 53%, consequently. To better understand the role of HA in the

198

PS/Fe(II)/HA system, different dosages of HA were added into the PS/Fe(II) system at 2 hour

199

intervals. As it could be observed that when the molar ratios of HA-to-Fe(II) increased from 0.5/1

200

to 4/1, the removal efficiency of BDE209 was dramatically increased compared to the control

201

process, and the removal efficiency increased from 53% to 66%. However, as the dosage of HA

202

further augmented, the degradation rates of BDE209 barely increased. And when the molar ratios of

203

HA-to-Fe(II) added from 4/1 to 8/1, the removal rate of BDE209 decreased from 66% to 65%,

204

unfortunately. It was clear that although high levels of NH3 OH+ could promote the transformation

205

of Fe(III) to Fe(II), when the contents of NH3OH+ further increased, it would act as a scavenger for

206

SO4−• and HO• with high reaction rates, i.e., k < 5.0×108 M-1 s-1 and k = 1.5×107 M-1 s-1 for HO•[27]

207

and SO4−•[33], which led to the decline of BDE209 removal efficiency. Consequently, the molar

208

ratio of HA/Fe(II) = 4/1 was the optimal HA dosage.

209 10

(6)

210

3.5. BDE209 oxidation at different initial BDE209 concentrations

211

The effects of various initial BDE209 levels (5, 10 and 20 mg kg-1 ) on the BDE209 removal

212

were investigated at a persulfate dosage of 0.2 M, an initial pH of 3.0 and a fixed molar ratio of

213

Fe(II)/PS at 0.5/1. As demonstrated in Fig. 5, the removal efficiency of BDE209 generally

214

decreased with the increasing BDE209 concentrations. The removal rates dropped from 65% to

215

32% as the initial BDE209 levels increased from 5 to 20 mg kg-1, and the rate constants decreased

216

from 0.57 to 0.14 h-1, correspondingly. As the initial BDE209 dosage increased, it would result in

217

the contact between BDE209 with oxidants more frequently, leading to more BDE209 being

218

oxidized. The absolute quantity of oxidized BDE209 molecules raised from 1.6×10-2 to 3.2×10-2 mg

219

as the initial BDE209 contents added from 5 to 20 mg kg-1, respectively.

220 221

3.6. Degradation intermediates and pathways of BDE209

222

The possible debrominated products of BDE209 were determined by the retention time in a

223

NCI mode, recording the m/z 79 and 81 as major ions[34]. And other intermediates were identified

224

in an EI mode. Based on the total ion current (TIC) of GC-MS analysis, about 9 intermediates were

225

detected.

226

Using a NCI mode, five debrominated products including BDE99, BDE183, BDE203,

227

BDE206 and BDE207 were determined (presented in Fig. S1); when using an EI mode, another four

228

products (presented in Fig. S2) including

229

3,5-dibromo-6-(2,4-dibromo-3,5-dihydroxyphenoxy)benzene-1,2,4-triol,

230

3,5-dibromo-2-(2,4-dibromo-5-hydroxyphenoxy)phenol,

231

5-bromo-2-(2,4,5-tribromophenoxy)phenol and 4,6-dibromobenzene-1,3-diol were identified. The 11

232

intermediates determination was preceded as follows: The mass fragment ions at m/z 563.6 for

233

[M-2H], 281.0 for [C6HBr2O3]−, 171 for [C6H3BrO], which may be labeled as

234

3,5-dibromo-6-(2,4-dibromo-3,5-dihydroxyphenoxy)benzene-1,2,4-triol or their isomers

235

(C12H6Br4O6, Mr. 566); The mass fragment ions at m/z 516.5, 403.0, 355.1, 246.1 were consistent

236

with [M-H], [M-HBr-2OH], [M-2Br-3H] and [M-3Br-2OH], which may be labeled as

237

3,5-dibromo-2-(2,4-dibromo-5-hydroxyphenoxy)phenol or their isomers (C12H6Br4O3, Mr. 518);

238

Additionally, the mass fragment ions at m/z 503.1, 341.0, 326.9, 173.1 and 91.1 may be labeled as

239

5-bromo-2-(2,4,5-tribromophenoxy)phenol or their isomers (C12H6Br4O2, Mr. 502); The major

240

fragment ions at m/z 270.4, 252.1, 191.2, 108.2, 91.2 and 79.2 may be labeled as

241

4,6-dibromobenzene-1,3-diol or their isomers (C6H4Br2O2, Mr. 270).

242

According to the analyzed results by GC-MS, the potential degradation pathways of BDE209

243

by ferrous-activated persulfate oxidation were proposed in Fig. 6. In general, the degradation of

244

BDE209 could be divided into four steps. Firstly, BDE209 transformed to the low brominated

245

intermediates including BDE99, BDE183, BDE203, BDE206 and BDE207 by debromination; Then,

246

HO• successively attacked them by nucleophilic substitution of bromine and the addition reactions,

247

yielding 3,5-dibromo-6-(2,4-dibromo-3,5-dihydroxyphenoxy)benzene-1,2,4-triol,

248

3,5-dibromo-2-(2,4-dibromo-5-hydroxyphenoxy)phenol,

249

5-bromo-2-(2,4,5-tribromophenoxy)phenol and 4,6-dibromobenzene-1,3-diol or their isomers;

250

Thirdly, free radicals attacked the ether bond, yielding 4,6-dibromobenzene-1,3-diol or their

251

isomers; Finally, the degradation of BDE209 was ended by the mineralization to CO2, Br- and

252

H2O.

253 12

254

4. Conclusions

255

In the present study, a sulfate radical-based advanced oxidation process was conducted for

256

BDE209 removal in spiked soil. Various factors were determined in a Fe(II)/PS system, such as pH,

257

BDE209 and ferrous contents. The results indicated that a 0.5/1 molar ratio of Fe(II)/PS was

258

observed to be best. The optimal pH and PS concentration for maximum degradation of BDE209

259

were 3.0 and 0.2 mol L-1, respectively. In addition, increasing concentrations of sodium persulfate

260

and ferrous ions appropriately contributed to the degradation of BDE209 within the applied

261

Fe(II)/PS molar ratios. HA was also used in the PS/Fe(II) system and further promoted the

262

degradation efficiency of BDE209. As the ratio of HA/Fe(II) increased from 0/1 to 4/1, the removal

263

efficiency were improved from 53% to its highest value of 66%.

264 265

Acknowledgements

266

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

267

(41371467), the Shanghai Pujiang Program (15PJD013), and the National Key Research and

268

Development Program (2016YFD0800405).

269 270 271 272 273 274

References [1] C.A.D. Wit, An overview of brominated flame retardants in the environment, Chemosphere, 46 (2002) 583-624. [2] W. Zhen, G. Na, X. Ma, L. Ge, Z. Lin, Z. Yao, Characterizing the distribution of selected PBDEs in soil, moss and reindeer dung at Ny-Ålesund of the Arctic, Chemosphere, 137 (2015) 9-13.

13

275 276

[3] J. Puckett, S.V.T. Coalition, Scope, G. China, Exporting harm: the high-tech trashing of Asia, Diane Publishing Co., 2002.

277

[4] Y. Li, J. Li, L. Wang, Recycling of PBDEs Containing Plastics from Waste Electrical and

278

Electronic Equipment (WEEE): A Review, in: e-Business Engineering (ICEBE), 2013 IEEE 10th

279

International Conference on, 2013, pp. 407-412.

280

[5] Y. Luo, X.J. Luo, Z. Lin, S.J. Chen, J. Liu, B.X. Mai, Z.Y. Yang, Polybrominated diphenyl

281

ethers in road and farmland soils from an e-waste recycling region in Southern China: Concentrations,

282

source profiles, and potential dispersion and deposition, Science of the Total Environment, 407 (2009)

283

1105-1113.

284

[6] H. Aizhen, W. Nan, L. Ming, Z. Lihua, Z. Yingying, L. Zhifen, Y. Daqiang, T. Heqing,

285

Efficient oxidative debromination of decabromodiphenyl ether by TiO2-mediated photocatalysis in

286

aqueous environment, Environmental Science & Technology, 47 (2013) 518-525.

287 288

[7] K. Young-Soo, Q.X. Li, Reductive Debromination of polybrominated diphenyl ethers by zerovalent iron, Environmental Science & Technology, 39 (2005) 2280-2286.

289

[8] E.J. Kim, J.H. Kim, J.H. Kim, V. Bokare, Y.S. Chang, Predicting reductive debromination of

290

polybrominated diphenyl ethers by nanoscale zerovalent iron and its implications for environmental

291

risk assessment, Science of the Total Environment, s 470-471 (2014) 1553-1557.

292

[9] Y.F. Huang, Y.H. Huang, Identification of produced powerful radicals involved in the

293

mineralization of bisphenol A using a novel UV-Na(2)S(2)O(8)/H(2)O(2)-Fe(II,III) two-stage oxidation

294

process, Journal of Hazardous Materials, 162 (2009) 1211-1216.

14

295

[10] L. Chenju, C.J. Bruell, M.C. Marley, K.L. Sperry, Persulfate oxidation for in situ remediation

296

of TCE. I. Activated by ferrous ion with and without a persulfate-thiosulfate redox couple,

297

Chemosphere, 55 (2004) 1213-1223.

298

[11] Y. Ji, C. Dong, D. Kong, J. Lu, Q. Zhou, Heat-activated persulfate oxidation of atrazine:

299

Implications for remediation of groundwater contaminated by herbicides, Chemical Engineering

300

Journal, 263 (2015) 45-54.

301 302 303 304 305 306

[12] A. Tsitonaki, B.F. Smets, P.L. Bjerg, Effects of heat-activated persulfate oxidation on soil microorganisms, Water Research, 42 (2008) 1013-1022. [13] C.W. Wang, C. Liang, Oxidative degradation of TMAH solution with UV persulfate activation, Chemical Engineering Journal, 254 (2014) 472-478. [14] L. Yating, L. Chenju, C. Jiunhua, Feasibility study of ultraviolet activated persulfate oxidation of phenol, Chemosphere, 82 (2011) 1168-1172.

307

[15] C.C. Lin, L.T. Lee, L.J. Hsu, Performance of UV/S2O82- process in degrading polyvinyl

308

alcohol in aqueous solutions, Journal of Photochemistry & Photobiology A Chemistry, 252 (2013) 1-7.

309

[16] O.S. Furman, A.L. Teel, R.J. Watts, Mechanism of base activation of persulfate,

310

Environmental Science & Technology, 44 (2010) 6423-6428.

311

[17] D.Y.S. Yan, I.M.C. Lo, Removal effectiveness and mechanisms of naphthalene and heavy

312

metals from artificially contaminated soil by iron chelate-activated persulfate, Environmental Pollution,

313

178 (2013) 15-22.

314

[18] C.H. Yen, K.F. Chen, C.M. Kao, S.H. Liang, T.Y. Chen, Application of persulfate to

315

remediate petroleum hydrocarbon-contaminated soil: Feasibility and comparison with common

316

oxidants, Journal of Hazardous Materials, 186 (2011) 2097-2102. 15

317

[19] A. Tsitonaki, B. Petri, M. Crimi, H. Mosbæk, R.L. Siegrist, P.L. Bjerg, In situ chemical

318

oxidation of contaminated soil and groundwater using persulfate: A Review, Environmental Science &

319

Technology, 40 (2010) 55-91.

320

[20] L. Zhou, W. Zheng, Y. Ji, J. Zhang, C. Zeng, Y. Zhang, Q. Wang, X. Yang, Ferrous-activated

321

persulfate oxidation of arsenic(III) and diuron in aquatic system, Journal of Hazardous Materials, 263

322

(2013) 422-430.

323

[21] H. Kusic, I. Peternel, S. Ukic, N. Koprivanac, T. Bolanca, S. Papic, A.L. Bozic, Modeling of

324

iron activated persulfate oxidation treating reactive azo dye in water matrix, Chemical Engineering

325

Journal, 172 (2011) 109-121.

326

[22] E. Viollier, P.W. Inglett, K. Hunter, A.N. Roychoudhury, P.V. Cappellen, The ferrozine

327

method revisited: Fe(II)/Fe(III) determination in natural waters, Applied Geochemistry, 15 (2000)

328

785-790.

329 330 331 332 333 334 335 336

[23] L. Johan, M. Gábor, Kinetic and thermodynamic properties of the aminoxyl (NH2O*) radical, Journal of Physical Chemistry A, 110 (2006) 192-197. [24] G. Bengtsson, S. Fronæus, L. Bengtsson-Kloo, The kinetics and mechanism of oxidation of hydroxylamine by iron(III), J.Chem.Soc.Dalton Trans, 12 (2002) 2548-2552. [25] M.D. Johnson, B.J. Hornstein, The kinetics and mechanism of the ferrate(VI) oxidation of hydroxylamines, Inorganic Chemistry, 42 (2003) 6923-6928. [26] P. Neta, R.E. Huie, A.B. Ross, Rate constants for reactions of inorganic radicals in aqueous solution, Journal of Physical & Chemical Reference Data, 17 (1988) 1027-1284.

16

337

[27] G.V. Buxton, C.L. Greenstock, W.P. Helman, A.B. Ross, Critical review of rate constants for

338

reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals (OH/O− in aqueous solution,

339

Journal of Physical & Chemical Reference Data, 17 (1988) 513-886.

340 341 342 343

[28] K.F. Chen, S.H. Liang, Methyl tert-butyl ether (MTBE) degradation by ferrous ion-activated persulfate oxidation: feasibility and kinetics studies, Water Environment Research, 81 (2009) 687-694. [29] G.R. Peyton, The free-radical chemistry of persulfate-based total organic carbon analyzers, Marine Chemistry, 41 (1993) 91-103.

344

[30] W. Chu, Y.R. Wang, H.F. Leung, Synergy of sulfate and hydroxyl radicals in

345

UV/S2O82−/H2O 2 oxidation of iodinated X-ray contrast medium iopromide, Chemical Engineering

346

Journal, 178 (2011) 154-160.

347

[31] J. Deng, Y. Shao, N. Gao, Y. Deng, S. Zhou, X. Hu, Thermally activated persulfate (TAP)

348

oxidation of antiepileptic drug carbamazepine in water, Chemical Engineering Journal, 228 (2013)

349

765-771.

350

[32] Z. Jing, M. Jun, C. Liwei, L. Xuchun, G. Yinghong, X. Pengchao, P. Chao, Rapid

351

acceleration of ferrous iron/peroxymonosulfate oxidation of organic pollutants by promoting

352

Fe(III)/Fe(II) cycle with hydroxylamine, Environmental Science & Technology, 47 (2013)

353

11685-11691.

354

[33] A. Rastogi, S.R. Al-Abed, D.D. Dionysiou, Effect of inorganic, synthetic and naturally

355

occurring chelating agents on Fe(II) mediated advanced oxidation of chlorophenols, Water Research,

356

43 (2009) 684-694.

17

357

[34] M.J. La Guardia, R.C. Hale, E. Harvey, Detailed polybrominated diphenyl ether (PBDE)

358

congener composition of the widely used penta-, octa-, and deca-PBDE technical flame-retardant

359

mixtures, Environmental Science & Technology, 40 (2006) 6247-6254.

360

18

361

Table captions:

362

Table1 Pseudo-first-order degradation rate constants of BDE209 at different operating conditions.

19

363

Table 1 Pseudo-first-order degradation rate constants of BDE209 at different operating conditions. pH

kBDE209 (h-1)

Half-life (h)

R2

3.0

0.48

1.44

0.96

5.0

0.44

1.58

0.98

7.0

0.41

1.69

0.97

9.0

0.34

2.04

0.93

364

20

365

Figure captions:

366

Fig. 1. BDE209 oxidation at different ferrous ion concentrations.

367

Fig. 2. BDE209 oxidation at different persulfate concentrations.

368

Fig. 3. BDE209 oxidation at different initial pH values.

369

Fig. 4. BDE209 oxidation at different hydroxylamine dosages.

370

Fig. 5. BDE209 oxidation at different initial BDE209 concentrations.

371

Fig. 6. Proposed oxidative degradation pathways of BDE209.

372

Support information. Mass spectrograms of the intermediate products.

21

1.0

0.8

Ct/C0

0.6

Control

0.4

2+

Fe /PS=0.1/1 2+

Fe /PS=0.2/1 2+

Fe /PS=0.5/1

0.2

2+

Fe /PS=1/1 2+

Fe /PS=2/1 0.0 0

373

1

2

3

4

5

6

Reaction time/h

374

Fig. 1. BDE209 oxidation at different ferrous ion concentrations. Experimental conditions:

375

[BDE209] = 10 mg kg-1; pH = 3.0; [PS] = 0.2 M; [Fe2+] = 0-0.4 M.

22

1.0

0.8

Ct/C0

0.6

0.4

Control 0.01M 0.05M 0.1M 0.2M 1.0M

0.2

0.0 0

376

1

2

3

4

5

6

Reaction time/h

377

Fig. 2. BDE209 oxidation at different persulfate concentrations. Experimental conditions: pH =

378

3.0; [BDE209] = 10 mg kg-1; [Fe(II)]/[PS] = 0.5/1; [PS] = 0-1.0 M.

23

1.0

3.0 5.0 7.0 9.0

0.8

Ct/C0

0.6

0.4

0.2

0.0 0

379

1

2

3

4

5

6

Reaction Time/h

380

Fig. 3. BDE209 oxidation at different initial pH values. Experimental conditions: [BDE209] = 10

381

mg kg-1; [Fe(II)] = 0.1 M; [PS] = 0.2 M; pH0 = 3.0-9.0.

24

1.0 2+

HA/Fe =0/1 2+

HA/Fe =0.5/1 2+

0.8

HA/Fe =1/1

Adding different amount of HA

2+

HA/Fe =2/1 2+

HA/Fe =4/1 2+

HA/Fe =8/1

Ct/C0

0.6

0.4

0.2

0.0 0

382

1

2

3

4

5

6

Reaction time/h

383

Fig. 4. BDE209 oxidation at different hydroxylamine dosages. Experimental conditions: pH = 3.0;

384

[BDE209] = 10 mg kg-1; [Fe(II)] = 0.1 M; [PS] = 0.2 M; [HA] = 0-0.8 M.

385

25

1.0

-1

5 mg kg

-1

10 mg kg 0.8

-1

20 mg kg

Ct/C0

0.6

0.4

0.2

0.0 0

386

1

2

3

4

5

6

Reaction Time/h

387

Fig. 5. BDE209 oxidation at different initial BDE209 concentrations. Experimental conditions: pH

388

= 3.0; [BDE209] = 5-20 mg kg-1; [Fe(II)] = 0.1 M; [PS] = 0.2 M.

389

26

390 391

Fig. 6. Proposed oxidative degradation pathways of BDE209.

27

392

Supporting information

393 394 395 396 397 398 399 400

Fig. S1. GC-NCI-MS chromatography spectrums of BDE209 degradation intermediates (A) and

401

PBDEs standard substance (B).

402

28

403 404 405 406 407 408 409 410 411 412 413 414 415 416 417 418 419 420

Fig. S2. Proposed structures and mass spectrograms of BDE209 degradation products.

421

29

422

Highlights

423 424 425 426 427 428 429 430

(1) Effective oxidation of BDE209 was achieved by Fe(II)-activated sodium persulfate. (2) Initial persulfate and Fe2+ concentrations, pH and HA dosages affected BDE209 degradation. (3) HA can greatly facilitate BDE209 oxidation by reduction of ferric iron in a Fe(II)/PS system. (4) The BDE209 degradation pathways and intermediate products were proposed.

431

30