Synergistic effects of HSO5− in the gamma radiation driven process for the removal of chlorendic acid: A new alternative for water treatment

Accepted Manuscript Synergistic effects of HSO5 − in the gamma radiation driven process for the removal of chlorendic acid: A new alternative for water treatment Noor S. Shah, Javed Ali Khan, Ala’a H. Al-Muhtaseb, Murtaza Sayed, Behzad Murtaza, Hasan M. Khan PII: DOI: Reference:

S1385-8947(16)30977-9 http://dx.doi.org/10.1016/j.cej.2016.07.031 CEJ 15480

To appear in:

Chemical Engineering Journal

Received Date: Revised Date: Accepted Date:

19 March 2016 30 June 2016 7 July 2016

Please cite this article as: N.S. Shah, J.A. Khan, A.H. Al-Muhtaseb, M. Sayed, B. Murtaza, H.M. Khan, Synergistic effects of HSO5 − in the gamma radiation driven process for the removal of chlorendic acid: A new alternative for water treatment, Chemical Engineering Journal (2016), doi: http://dx.doi.org/10.1016/j.cej.2016.07.031

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1

Synergistic effects of HSO5− in the gamma radiation driven process for

2

the removal of chlorendic acid: A new alternative for water treatment

3 4

Noor S. Shah1, 2, *, Javed Ali Khan2, Ala'a H. Al-Muhtaseb3, Murtaza Sayed2, 4,

5

Behzad Murtaza1, Hasan M. Khan2, *

6 1

7

Department of Environmental Sciences, COMSATS Institute of Information

8 9

Technology, Vehari 61100, Pakistan 2

Radiation Chemistry Laboratory, National Centre of Excellence in Physical Chemistry,

10 11

University of Peshawar, Peshawar 25120, Pakistan 3

Petroleum and Chemical Engineering Department, Faculty of Engineering, Sultan

12 13

Qaboos University, Muscat-Oman 4

Department of Chemistry, COMSATS Institute of Information Technology,

14

Abbottabad 22060, Pakistan

15 16 17 18

Corresponding authors Noor S. Shah ([email protected], [email protected]) Hassan M. Khan ([email protected])

19 20

Tel: +9267-3001606

1

Abstract

21 22

Removal of chlorendic acid, an emerging water pollutant and potential

23

carcinogenic, was investigated by gamma radiation in the absence and presence of

24

peroxymonosulfate (PMS, HSO5−). The removal of chlorendic acid (1.40 µM initial

25

concentration) by gamma radiation was promoted with PMS, i.e., 95% compared to 82%

26

in the absence of PMS, at an absorbed dose of 1000 Gy. The removal of chlorendic acid

27

by gamma-ray/PMS process was due to •OH and SO4•−. Second-order rate constants of

28

5.90 × 109, 1.75 × 109, and 2.05 × 109 M−1 s−1 for chlorendic acid with eaq−, •OH, and

29

SO4•−, respectively, were determined. The removal efficiency of chlorendic acid was

30

promoted with increasing initial PMS concentration and decreasing initial target

31

contaminant concentration. The removal of chlorendic acid by gamma-ray/PMS was

32

inhibited in the presence of CO32−, NO2−, p-CBA, m-TA, and alcohols. The presence of

33

Fe2+, Cu+, and Fe3+ with gamma-ray/PMS promoted removal efficiency of chlorendic

34

acid from 78% to 99, 94, and 89%, respectively, at 592 Gy. The degradation of

35

chlorendic acid by •OH and SO4 •− was found to be initiated at the carboxylate group as

36

could be revealed from nature of the transformation by-products. Nevertheless, this study

37

concluded that gamma-ray/PMS is of practical importance in treatment of natural water

38

containing chlorendic acid, as potential detoxification of chlorendic acid solution can be

39

revealed from 83% loss of chloride ion at 3000 Gy. In addition, gamma-ray/PMS process

40

achieved efficient removal of chlorendic acid even in the presence of commonly found

41

inorganic ions in natural water.

42 43

Key words: AOTs; Chlorendic acid; Gamma radiation; PMS; Water treatment.

44 2

45 46

1. Introduction

47

Water pollution by xenobiotics is a widespread problem throughout the world and

48

is becoming worst with time due to increased population and consequently increased

49

industrialization and agricultural activities etc. The most important xenobiotics

50

contributing greatly into water pollution and of environmental concern are

51

organochlorine compounds (OCC), characterized by their greater persistency, non-

52

biodegradability, and high toxicity [1]. Due to their greater toxicity, most classes of OCC

53

have been banned, however, some are reported to be still synthesized and used on large

54

scale in different countries of the world [2]. Among these, one important and highly toxic

55

OCC

56

dicarboxylic acid) which is still synthesized and used commercially on large scale in

57

many countries, such as USA and Belgium [3]. The most important uses of chlorendic

58

acid include as a flame retardant in polyurethane foams, wool, resins, paints, piping, in

59

the synthesis of metal organic frameworks, and as an extreme pressure lubricant [4]. The

60

wide spread applications led to extensive use of chlorendic acid. It has been reported that

61

in the years from 1986 to 2002, in the USA, production and import of chlorendic acid

62

totaled 500,000 lb to 10 million pounds [5]. The US Environmental Protection Agency

63

(US EPA) classified chlorendic acid as the most toxic and highly persistent compound

64

[5]. Chlorendic acid can causes irritation in skin and eyes, lung cancer, and gene mutation

65

etc. [3, 6]. The wide range applications have resulted in increased contamination of water

66

resources with chlorendic acid [4]. Chlorendic acid also enters into the aquatic

67

environment as a degradation product of cyclodiene pesticides, such as endosulfan. Due

is

chlorendic

acid

(1,4,5,6,7,7-hexachlorobicyclo-(2,2,1)-hept-5-ene-2,3-

3

68

to frequent and large scale discharges into aquatic environment as well as greater

69

solubility and persistency in water, chlorendic acid has become a significant threat to

70

public and ecological health [4]. Although highly persistent and toxic, no guidelines are

71

suggested for control of chlorendic acid. The technologies, such as adsorption and

72

biodegradation have been reported inefficient in removal of chlorendic acid from water

73

environment [4]. So there is a need to develop cost effective, efficient, and

74

environmentally friendly technologies that can effectively remove chlorendic acid and

75

other organochlorine compounds from contaminated waters.

76

Recently advanced oxidation technologies (AOTs) that rely on in situ generation

77

of reactive radicals, such as hydroxyl radical have received considerable attention for

78

effective removal and mineralization of highly toxic, persistent, and recalcitrant emerging

79

water pollutants [7-11]. Hydroxyl radical (HR, •OH) with high redox potential of 2.72 V

80

(depending on the experimental conditions) has been reported to be highly reactive and

81

react effectively and non-selectively with organic compounds with a reported second-

82

order rate constant of 108-1010 M−1 s−1 [12]. Recently introduced sulfate radical (SR,

83

SO4•−), having redox potential of 2.5-3.1 V, have received significant attention due to

84

their wide contribution into the efficient removal of organic pollutants [12, 13]. On

85

contrary to •OH, SO4•− reacts selectively with most of the organic compounds with a high

86

second-order rate constant of 107-1010 M−1 s−1 [12]. Both •OH and SO4•− reacts with

87

organic compounds through three different ways, such as addition to unsaturated bond,

88

hydrogen abstraction, and electron abstraction from aromatic ring, double bond, and

89

saturated carbon or carboxylate group [9, 12]. As a result, in the present study, removal

90

of chlorendic acid from water environment was investigated by both •OH- and SO4•−-

4

91

AOTs. PMS (HSO5−), commonly known as Oxone® and an active component of a triple

92

salt, 2KHSO5•KHSO4•K2SO4, was used as a source of •OH and SO4•− through the

93

activation of gamma radiation and transition metal ions [12]. The high redox potential,

94

1.82 V, of PMS [14] makes its easier activation by electron from gamma radiation and

95

transition metal ions. The easy electron donating property as well as abundant

96

concentration in natural water make transition metal ions the best option for the

97

activation of PMS and is thus important for potential practical applications [11].

98

Similarly gamma radiations are predominant precursors of electron and can be effective

99

in the activation of PMS [15, 16]. Besides, gamma radiation has been proved to be the

100

most competent, efficient, and environmentally friendly treatment technologies and also

101

recommended by the international agencies for detoxification of pollutants [15, 16].

102

The main aim of the present study is to transform the eaq− into •OH and SO4•− by

103

combining PMS with gamma radiation. Although eaq− is highly reactive species for the

104

removal of halogenated organic compounds, the presence of inorganic ions, such as

105

nitrate and transition metal ions, efficient scavengers of eaq−, in natural water makes its

106

practical application impossible [15]. Nitrate ion has low reactivity towards •OH and

107

SO4•− (k < 1 × 105 M−1s−1) and therefore, the efficiency of •OH and SO4•− based AOTs are

108

expected to be independent of this ion. In addition, the presence of transition metal ions

109

could have a positive effect on the efficiency of gamma-ray/PMS process due to their

110

ability to activate PMS through electron transfer mechanism and, hence, dual activation

111

of PMS is possible. Moreover, using relatively higher concentration of PMS could

112

possibly reduce the interference of dissolved oxygen, which hinders the activity of eaq−

113

based remediation technologies but not of •OH and SO4•− based remediation technologies.

5

114

So, in the present study the yield of •OH and SO4•− from the activation of PMS by gamma

115

radiation and transition metals in the presence of gamma radiation for the treatment of

116

chlorendic acid has been assessed for the first time in the present study. Different radical

117

scavengers and competition kinetic studies were conducted to investigate the yield and

118

performance of •OH and SO4•− in the removal of chlorendic acid from aqueous solution.

119

The toxicity evaluation and main degradation pathways of the removal of chlorendic acid

120

by •OH and SO4•− was also examined.

121

2. Materials and methods

122

2.1.

123

Chlorendic acid, chlorendic anhydride and 1,3-cyclopentadiene, 1,2,3,4,5,5-

124

hexachloro (purity ≥ 99%) were purchased from Supelco (PA, USA). Oxone® (Sigma–

125

Aldrich) was used as the oxidant with gamma radiation in the present study. Other

126

chemicals, i.e., oxetane, m-toluic acid (m-TA), p-chlorobenzoic acid (p-CBA), 2-

127

chlorobenzoic acid (2-CBA), 2-butyl alcohol, iso-propyl alcohol, ethanol, methanol,

128

phosphoric acid, sodium sulfate (Na2SO4), cuprous chloride (CuCl), sodium nitrite

129

(NaNO2), potassium chloride (KCl), potassium carbonate (K2CO3), sodium acetate

130

(CH3CO2Na), ferrous sulfate (FeSO4. 7H2O), and ferric chloride (FeCl3. 6H2O) were

131

purchased from Scharlau. All the chemicals used in the present study were of high purity

132

and used as received. All the solutions were prepared in ultra pure water (resistivity, 18.2

133

MΩ.cm) obtained from Milli-Q® system (Millipore).

Materials

134

6

135 136

2.2.

137

The analysis of chlorendic acid, chlorendic anhydride, and 1,3-cyclopentadiene,

138

1,2,3,4,5,5-hexachloro was carried out using an Agilent 6890 series gas chromatography

139

(GC) equipped with Ni63 electron capture detector (ECD) and an HP-5 (5% phenyl

140

methylsiloxane) capillary column (30 m × 0.25 mm I.D., 0.25 µm particle size). Prior to

141

analysis with GC-ECD, the chlorendic acid, chlorendic anhydride, and 1,3-

142

cyclopentadiene, 1,2,3,4,5,5-hexachloro were extracted by solid phase extraction (SPE,

143

SUPELCO, PA, USA) technique using C18 cartridge. Ethanol was used as a solvent to

144

elute the adsorbed material on SPE cartridge for direct injecting into the injector of the

145

GC-ECD using liquid injector. The GC-ECD analysis was performed using injector, inlet,

146

and detector temperature of 250 oC, 220 oC, and 350 oC, respectively. The oven

147

temperature was programmed as; initially 80 oC (hold for 2 minutes), then 150 oC (hold

148

for 2 minutes) at a rate of 20 oC/min and finally increased to 220 oC (hold for 10 minutes)

149

at 10 oC/min rate. Nitrogen gas with a purity of 99.99% was used as a carrier gas at a

150

flow rate of 1.0 mL/min.

Analysis

151

The p-CBA and m-TA were quantified with an Agilent 1200 series high

152

performance liquid chromatography (HPLC, Agilent) using Eclipse XDB-C18 column

153

(150 m × 4.6 mm, 5 µm particle size) and a variable wavelength detector (VWD) set at

154

234 nm. Methanol and water (10.0 mM phosphoric acid) at a ratio of 50:50 (v/v) was

155

used as an eluent at a flow rate of 1.0 mL/min.

156 157

Spectrophotometric measurement was carried out to quantify the concentration of PMS using a UV-vis spectrophotometer (PerkinElmer, Lambda 650) [12].

7

158

The analysis of chloride and acetate ions formed from the decomposition of

159

chlorendic acid was done by ion chromatography (IC, Metrohm) using electrical

160

conductivity detector. The IC method for anion determination using Assup-5 column

161

(250/4.0 mm) and 3.2 mM Na2CO3/1mM NaHCO3 as an eluent at a flow rate of 0.75

162

mL/min was used.

163

The transformation by-products of chlorendic acid, both organic and inorganic,

164

were analyzed based on comparison of their retention times with those of authentic

165

standard under the similar experimental conditions [15, 17].

166

2.3.

167

Gamma-ray from Cobalt-60 source (model Issledovadel, origion USSR), installed

168

at the Nuclear Institute for Food and Agriculture (NIFA), Tarnab, Peshawar, Pakistan

169

was used for treatment of aqueous chlorendic acid solution. The calibration of the source

170

for the dose rate determination is explained in our previous study with a typical dose rate

171

calculated to be 296 Gy/h [18].

Gamma irradiation source and procedure

172

An aqueous solution of chlorendic acid was irradiated with gamma radiation for

173

different absorbed doses from 0−1000 Gy corresponding to irradiation time of 0−3.38 h.

174

However, for analysis of chloride ion, chlorendic acid aqueous solution was irradiated for

175

0-3000 Gy (irradiation time 0−10.13 h). The irradiation treatment of chlorendic acid was

176

conducted in 17 mL air tight Pyrex glass tubes at a normal pH, i.e., pH 5.8 of Milli-Q

177

water. However, with the addition of PMS to the aqueous solution of chlorendic acid, the

178

pH decreased from 5.8 to 5.2, due to the slight acidic nature of the PMS [14]. An aqueous

179

solution, 15 mL, of chlorendic acid was sparged with N2 gas for 30 minutes in air tight

180

test tubes and then put the stoppered test tubes in the gamma-ray sources for irradiation

8

181

treatment for predetermined time. The experiments were performed in triplicates and the

182

standard error of the mean is shown as the error bars in figures.

183

When treated with gamma radiation, radiolysis of dilute aqueous solution yield

184

reactive species as shown in Eq. (1) with their radiation chemical yield (G-value), defined

185

as the number of species (i.e., atoms, ions, and molecules) in micromole (µmol) formed

186

or consumed by absorption of 1 joule of radiation energy, in small bracket [15].

187

H2O --^-^-^--> •OH (0.29), •H (0.06), eaq− (0.28), H2 (0.047), H2O2 (0.07), H3O+ (0.27)

188

(1)

189

The hydroxyl radicals (•OH), hydrogen atom (•H) and aqueous electron (eaq−) are the

190

most reactive species due to their greater yield and high redox potential [15, 16]. When

191

combined with PMS, eaq− from gamma radiation react with PMS and yield •OH and SO4•−

192

as shown by reactions (1) and (2) in Table 1 [19, 20].

193

3. Results and Discussion

194

3.1.

195

The radiolytic degradation of chlorendic acid was assessed in the absence

196

(control) and presence of PMS to reveal the performance of PMS in the removal of

197

chlorendic acid in the presence of gamma radiation. When treated with gamma radiation,

198

the •OH, •H, and eaq− produced as shown in Eq. (1) are reported to contribute into

199

degradation of the target contaminant [15, 16]. Table 2 shows that at an absorbed dose of

200

1000 Gy, gamma radiation (control) resulted in 82% removal of chlorendic acid. In the

201

presence of PMS, an efficient scavenger of eaq−, instead of inhibition, however, removal

202

efficiency of chlorendic acid was promoted to 95%. The elevation of removal efficiency

203

of chlorendic acid in the presence of PMS suggests significant performance of PMS in

Degradation kinetic of chlorendic acid with the presence of PMS

9

204

the presence of gamma radiation. The possible reason for this looked to be that eaq− from

205

gamma radiation decomposes PMS due to easier eaq− accepting property of PMS and

206

yield •OH and SO4•− as shown by reactions (1) and (2) in Table 1 [19, 20]. As a result,

207

concentration of •OH is increased. Both •OH and SO4•− are highly reactive and therefore

208

readily attack the target contaminant [12]. The higher removal efficiency of chlorendic

209

acid by gamma-ray/PMS process possibly is due to increase in the yield of •OH and

210

significant contribution of •OH and SO4•− in the removal of chlorendic acid.

211

To assess formation of •OH and SO4•− from the reactions of PMS in the presence

212

of gamma radiation as well as involvement of •OH and SO4•− in the removal of chlorendic

213

acid, competition kinetic study was carried out using para-chlorobenzoic acid (p-CBA)

214

and meta-toluic acid (m-TA) as radical scavenger for •OH and SO4•−, respectively. The

215

faster kinetic of p-CBA with •OH and m-TA with SO4•− (as shown by reactions (3) and

216

(4) in Table 1) might led to significant competition of p-CBA and m-TA with chlorendic

217

acid for •OH and SO4•− and consequently lowering in removal efficiency of chlorendic

218

acid is expected [12]. At an absorbed dose of 1000 Gy, removal efficiency of chlorendic

219

acid by gamma-ray/PMS process was 45 and 40% in the presence of p-CBA and m-TA,

220

respectively, as compared to 95% in the absence of either scavenger (Table 2). The

221

higher inhibition in removal efficiency of chlorendic acid in the presence of both p-CBA

222

and m-TA suggests formation of •OH and SO4•− from reactions of PMS in the presence of

223

gamma radiation and also involvement of •OH and SO4•− in the degradation of chlorendic

224

acid. At an absorbed dose of 1000 Gy, 80% of PMS was decomposed (data not shown)

225

that further verify yield of •OH and SO4•− from reactions of PMS in the presence of

226

gamma radiation. The removal efficiency of chlorendic acid was promoted with

10

227

increasing absorbed gamma-ray dose both in the presence and absence PMS which was

228

consistent to the finding in previous studies [15, 16, 18].

229

The radiolytic degradation of chlorendic acid followed pseudo-first-order kinetic

230

and observed pseudo-first-order degradation rate constants (kobs, Gy−1) were determined

231

using integrated pseudo-first-order rate equation (Eq. (2)) [15].

232

 C  − ln   = k o bs D  C0 

233

Table 2 shows that kobs for radiolytic degradation of chlorendic acid was higher in the

234

presence of PMS and gamma radiation process only and least for the gamma radiation

235

with PMS process in the presence of both p-CBA and m-TA. The kobs were used to

236

calculate the dose required for 90% removal of chlorendic acid (i.e., D0.9) under the

237

studied conditions using the following equation (Eq. (3)).

238

D 0.90 =

239

The D0.9 was 794, 1350, 3710, and 4260 Gy for gamma-ray/PMS process, gamma-ray

240

only (control), gamma-ray/PMS/p-CBA, and gamma-ray/PMS/m-TA, respectively (Table

241

2). The G-values for radiolytic degradation of chlorendic acid at different absorbed doses

242

(i.e., 148−1000 Gy) were calculated using Eq. (4) [15].

243

ln 10 k obs

G-value=

[R] 1.0 × 106 µmol/J D

(2)

(3)

(4)

244

The [R] in Eq. (4) refer to change in chlorendic acid concentration in µM at a given dose

245

while D is for absorbed dose in Gy. The G-value for radiolytic degradation of chlorendic

246

acid at the studied experimental conditions, i.e., control, PMS only, and PMS with p-

247

CBA and with m-TA was higher at lower absorbed dose than at higher absorbed dose and

248

was consistent with the finding in previous studies [15, 21]. The possible reason for the 11

249

observed trend looked to be the increased competition between parent compound and its

250

by-products for reactive radicals at higher absorbed dose [15]. The G-value for the

251

degration of chlorendic acid was the highest for gamma-ray/PMS followed by control and

252

least for gamma-ray/PMS/p-CBA and gamma-ray/PMS/m-TA (Table 2).

253

3.2.

Determination of second-order rate constant of chlorendic acid with •OH

and SO4•−

254 255

Since •OH and SO4•− that react with chlorendic acid results from the reaction of

256

eaq− with PMS, the eaq− instead of reacting with PMS might react with chlorendic acid and

257

could influence the reactivity of eaq− with PMS and so the yield of •OH and SO4 •−. To

258

know reactivity of eaq− with chlorendic acid, second-order rate constant of eaq− with

259

chlorendic acid ( keaq− /chlorendic acid ) was determined using gamma radiation in the absence

260

of PMS following the method reported by Vel Leitner et al. [22] as shown in Eq. (5). This

261

was done in the presence of 2-butyl alcohol (2-BuOH) used as a scavenger for •OH and

262

•

263

chlorophenol (2-CP) as a competitor for eaq− [15, 23].

264

ln[

H due to their high second-order rate constants (reactions (5) and (6) in Table 1) and 2-

ke − /Chlorendic acid Chlorendic acidD 2 − CPD ] = aq ln[ ] Chlorendic acid0 ke − /2−CP 2 − CP0

(5)

aq

265

The ln[ Chlorendic acid D ] and ln[ 2 − C PD ] refer to ln of the change in concentration of C hlorendic acid 0

2 − CP0

266

chlorendic acid and 2-CP at different absorbed doses (D) with respect to the initial

267

concentration while ke

− aq

/2 − CP

and ke

− aq

/ Chlorendic acid

refer to the second-order rate constants

268

of 2-CP and chlorendic acid with eaq−, respectively. Irradiating an aqueous solution

269

containing both chlorendic acid and 2-CP (i.e., 5 µM) for absorbed doses from 0−1000 12

270

Gy and putting the values in Eq. 5, the second-order rate constant of eaq− with chlorendic

271

acid was found to be 5.90 × 109 M−1 s−1, slightly smaller than second-order rate constant

272

of eaq− with PMS (i.e., 8.40 × 109 M−1 s−1).

273

The reactivity of eaq− with PMS, e.g., r( eaq − / PMS )) and with chlorendic acid, i.e.,

274

r( eaq − / chlorendic acid ) depends on the initial concentration of PMS and chlorendic acid

275

as well as their second-order rate constant with eaq− as can be revealed from Eqs. (6) and

276

(7).

277

r( eaq − / PMS ) =

278

r( eaq − / chlorendic acid ) =

279

Due to small difference between second-order rate constants of eaq− with PMS and

280

chlorendic, several fold high concentration of PMS (i.e., [PMS] = 1.0 mM) than

281

chlorendic acid, i.e., [chlorendic acid] = 5.0 µM was used in gamma-ray/PMS process so

282

that almost all of eaq− from gamma-ray react with PMS and give considerable yield of

283

•

284

8400000 s−1) was found to be 285 times higher than reactivity of eaq− with chlorendic acid

285

(i.e., 29500 s−1) under the studied conditions. The greater reactivity of eaq− with PMS

286

attribute to significant yield of •OH and SO4•− that quickly react with the target

287

contaminant.

ke

− aq / PMS

[PMS]

(6)

ke − /chlorendic acid [chlorendic acid]

(7)

aq

OH and SO4•−. Putting the values in Eqs. (6) and (7), reactivity of eaq− with PMS (i.e.,

288

To assess reactivity of •OH and SO4•− with chlorendic acid, second-order rate

289

constants of chlorendic acid with •OH ( k • OH / chlorendic acid ) and with SO4•− ( kSO •− / chlorendic acid ) 4

290

were determined following Eq. (5) and using m-TA and p-CBA as a competitor for SO4•−

291

and •OH, respectively [23, 24]. For determining second-order rate constants of chlorendic 13

292

acid with SO4 •−, 2-BuOH was used with gamma-ray/PMS/m-TA as a scavenger for •OH

293

and •H due to their fast reactions (reactions (5) and (6) in Table 1) while for second-order

294

rate constant determination of •OH with chlorendic acid, oxetane and 2-chlorobenzoic

295

acid was used with gamma-ray/PMS/p-CBA as a scavenger for SO4•− and •H,

296

respectively, due to their fast reactions as shown in reactions (7) and (8) in Table 1 [23].

297

The same initial concentration of chlorendic acid as well as p-CBA and m-TA, i.e., 5.0

298

µM, was used and irradiated for different gamma-ray doses from 0−1000 Gy. The

299

second-order rate constants of 1.75 × 109 M−1 s−1 and 2.05 × 109 M−1 s−1 for chlorendic

300

acid with •OH and SO4•−, respectively, were subsequently obtained. Despite the high second-order rate constant of chlorendic acid with eaq− than with

301 302

•

OH and SO4•−, high removal efficiency of chlorendic acid, i.e., 95% was obtained in

303

gamma-ray/PMS process than 82% in gamma-ray process only (control) as already

304

discussed in section 3.1. Due to high redox potential and un-stability in aqueous solution,

305

PMS has been reported to cause autoxidation of the target contaminant and could be one

306

of the possible reasons for high removal efficiency in gamma-ray/PMS than control [12,

307

20].

308

According to Myzek et al. [25], removal efficiency of compound is not always

309

directly related to the second-order rate constant of the target compound with the reactive

310

species. It is also significantly influenced by the nature of the by-products formed

311

through the reaction of the given reactive species. If the by-products are more reactive

312

towards given reactive species than the parent compound, the removal efficiency of the

313

parent compound will be low and vice versa [25]. In the present study possibly •OH/SO4•−

314

are producing less reactive by-products of chlorendic acid than those produced by

14

315

aqueous electron, as a result, •OH/SO4•− gave higher removal efficiency than aqueous

316

electron.

317

It has been reported that despite de-aeration of the target compound solution with

318

nitrogen gas, it is expected that some oxygen still remained in the solution which

319

scavenged aqueous electron and converted it into less reactive superoxide radical anions

320

(O2•−) [23, 25]. However, in the presence of PMS, aqueous electron react with PMS

321

rather than oxygen and led to the formation of •OH and SO4•−. This could be another

322

possible reason for high removal efficiency of chlorendic acid in gamma-ray/PMS

323

process than in control.

324

3.3.

Effects of initial concentrations of PMS

325

In order to investigate the performance of PMS under gamma radiation, a detail

326

study of chlorendic acid (i.e., 1.40 µM) degradation was carried out at different initial

327

concentrations of PMS, i.e., 35.0, 70.0, 140.0 and 280.0 µM. The increased initial

328

concentrations of PMS led to faster removal of chlorendic acid under gamma radiation,

329

showing a removal efficiency of 85% at 1000 Gy for an initial PMS concentration of 35.0

330

µM as compared to 92% at 592 Gy for an initial PMS concentration of 280.0 µM (Figure

331

1). These findings were consistent to those in the previous studies [11-13]. In the

332

presence of gamma radiation, PMS yield •OH and SO4•− as shown by reactions (1) and

333

(2) in Table 1. Elevation of initial concentrations of PMS can thus probably led to

334

increase in the rate of formation of •OH and SO4•− and subsequently higher removal

335

efficiency of chlorendic acid is thus achieved. This study further suggests significant

336

contribution of PMS in the radiolytic degradation of chlorendic acid in our study.

337

15

338 339

3.4.

Effect of initial concentrations of chlorendic acid

340

Initial concentration of organic pollutants in natural aquatic environments is one

341

of the main factors affecting the efficiency of a remediation treatment technology. In

342

order to assess the aforementioned factor, the effect of initial concentrations of chlorendic

343

acid on its removal efficiency by gamma-ray/PMS was investigated in the present study

344

(Table 3). The absorbed gamma-ray dose was not proportional with initial concentrations

345

of chlorendic acid and at the same absorbed dose, removal efficiency of chlorendic acid

346

was found to be higher at lower initial concentration than at higher initial concentration.

347

These results were consistent to the findings in previous studies [15-18]. At an absorbed

348

dose of 592 Gy, the removal efficiency of chlorendic acid decreased from 96 to 61% with

349

increasing initial concentrations from 0.35 to 2.80 µM, respectively (Table 3). At each

350

studied concentration of chlorendic acid, kobs was calculated using Eq. (2) and was found

351

to decrease from 5.30 × 10−3 to 1.90 × 10−3 Gy−1 with increasing initial concentrations of

352

chlorendic acid from 0.35 to 2.80 µM, respectively (Table 3). The relationship between

353

initial concentrations of chlorendic acid and kobs can be expressed as a power function as

354

shown in Eq. (8)

355

kobs = 0.0032C0−0.472

R2 = 0.978

(8)

356

This decrease in removal efficiency (%) and kobs with increasing initial

357

concentrations of chlorendic acid is attributed to the decrease in ratio of reactive radicals

358

to the target contaminant and increased competition between the target contaminant and

359

transformation products for reactive radicals at higher initial concentration than at lower

360

initial concentration of chlorendic acid [11, 15]. The initial degradation rate of chlorendic

16

361

acid (µM/Gy) was calculated for the initial 296 Gy and was found to increase from 6.0 ×

362

10−4 to 3.4 × 10−3 µM/Gy when initial concentration of chlorendic acid was elevated from

363

0.35 to 2.80 µM (Table 3). The G-value (µmol/J) for radiolytic degradation of chlorendic

364

acid in the presence of PMS was also calculated at each studied concentration using Eq.

365

(4). Table 3 shows that G-value increased with increasing initial concentration of

366

chlorendic acid and was found to be 1.5 × 10−3 and 5.4 × 10−3 µmol/J for an initial

367

chlorendic acid concentration of 0.35 and 2.80 µM, respectively. The increase in initial

368

degradation rate and G-value with increasing initial concentrations of chlorendic acid

369

looked to be due to increase number of molecules of the target contaminant exposed to

370

reactive radicals at higher initial concentration than at lower initial concentration [11, 15].

371

The study of the effect of initial concentrations on degradation is beneficial to

372

suggest a guideline dose for treatment of chlorendic acid in contaminated water at

373

different concentrations.

374

3.5.

Influence of common inorganic anions on degradation of chlorendic acid

375

Inorganic anions are common constituents of natural water, found at different

376

concentrations ranging from 0.01 to 1.0 mM [12]. These anions have high second-order

377

rate constants with •OH and SO4•− as shown in reactions (9) − (16) in Table 1 and might

378

influence their reactivity with chlorendic acid. Therefore, investing the effects of these

379

anions was important for potential practical applications. The influence of these anions,

380

i.e., carbonate (CO32−), nitrite (NO2−), chloride (Cl−), and sulfate (SO42−) were

381

investigated on the radiolytic degradation of chlorendic acid in the presence of PMS

382

(gamma-ray/PMS). At an absorbed dose of 1000 Gy, removal efficiency of chlorendic

383

acid was 70, 89, 52, and 35% in the presence of Cl−, SO42−, CO32−, and NO2−,

17

384

respectively, as compared to 95% in the absence of either anion (Figure 2). The high

385

second-order rate constants and consequently greater reactivity of NO2− and CO32− with

386

•

387

SO4•− and consequently lowered removal efficiency of chlorendic acid in the presence of

388

these anions. These results were consistent to the finding in previous studies [12].

389

Besides greater scavenging of •OH and SO4 •−, the NO2− has been reported to significantly

390

scavenge eaq− that could also influence removal efficiency of chlorendic acid in the

391

presence of this anion [15, 23]. The lower second-order rate constant of SO42− with •OH

392

and SO4•− (reactions (13) and (14) in Table 1) might led to less scavenging of •OH and

393

SO4•− and consequently removal efficiency was inhibited to a smaller extent. Besides,

394

scavenging of •OH and SO4•− by SO42− has been reported to yield SO4•− and highly

395

reactive reducing reactive species, eaq−, respectively, (reactions (13) and (14) in Table 1)

396

[23] that might influence the removal of chlorendic acid. The removal efficiency of

397

chlorendic acid by gamma-ray/PMS with Cl− was also inhibited, however, to a lower

398

extent than by CO32−and NO2−, despite the high second-order rate constants of Cl− with

399

•

400

trend looked to be the formation of reactive •OH, Cl• and Cl•− through stepwise reactions

401

(17) − (19) as shown in Table 1 as well as the activation of PMS by Cl− yielding reactive

402

HOCl (reaction (20) in Table 1) [12, 14].

OH and SO4•− (reaction (9) – (12) in Table 1) might led to greater scavenging of •OH and

OH and SO4•− (reactions (15) and (16) in Table 1). The possible reason for the observed

403

3.6.

Influence of transition metal ions on degradation of chlorendic acid

404

Several studies reported the analysis of transition metals in natural water,

405

investigating the influence of these metals on the removal efficiency of chlorendic acid

406

by gamma radiation in the presence of PMS is therefore vital for potential practical

18

407

applications [26]. The removal efficiency of chlorendic acid by gamma-ray/PMS was

408

investigated in the present study in the presence of the most common transition metal

409

ions, i.e., cuprous (Cu+), ferrous (Fe2+) and ferric (Fe3+). The greater reactivity of these

410

metal ions with eaq− might inhibit performance of gamma-ray process only due to greater

411

scavenging of eaq− by these ions; however, the gamma-ray/PMS process was promoted in

412

the presence of these metal ions [15, 23]. On contrary to 95% removal of chlorendic acid

413

at 1000 Gy by gamma-ray/PMS process, the presence of Fe2+ and Cu+ with gamma-

414

ray/PMS resulted in 99% and 94% removal of chlorendic acid at 592 Gy, respectively

415

(Figure 3). The faster degradation of chlorendic acid in the presence of Fe2+ and Cu+ with

416

gamma-ray/PMS suggests increased activation of PMS and consequently increased the

417

rate of •OH and SO4•− formation. Due to high redox potential and greater electron

418

accepting property, PMS can be easily activated by transition metals catalysts, i.e., Fe2+

419

and Cu+ through transfer of electrons from metals to the central atom of PMS (as shown

420

in reactions (21) − (24) in Table 1) [11]. The slightly higher degradation of chlorendic

421

acid by Fe2+ than Cu+ looked to be due to faster kinetics of PMS with Fe2+ than with Cu+

422

(reactions (21) − (24) in Table 1). The dual activation of PMS by gamma radiation as

423

well as by Fe2+ and Cu+ in gamma-ray/PMS in the presence of Fe2+ and Cu+ possibly

424

increased the rate of formation of •OH and SO4•− and consequently resulted in higher

425

removal efficiency of chlorendic acid in our study. The removal of chlorendic acid was

426

faster in the presence of Fe3+ by gamma-ray/PMS process as well. At an absorbed dose of

427

1000 Gy, 99.3% removal of chlorendic acid was achieved by gamma-ray/PMS/Fe3+

428

process as compared to 95% in gamma-ray/PMS (Figure 3). The lower removal

429

efficiency of chlorendic acid in gamma-ray/PMS/Fe3+ than by gamma-ray/PMS/Fe2+ or

19

430

gamma-ray/PMS/Cu+ could be due to lower activation of PMS by Fe3+. It has been

431

reported that activation of PMS by Fe3+ yield a less reactive SO5•− (reaction (25) in Table

432

1) and involve transfer of electrons from PMS towards Fe3+ which is a slow process [11].

433 434

3.7.

Confirmation of •OH and SO4•− formation from transition metals-

mediated PMS

435

The yield of •OH and SO4•− from activation of PMS by transition metals in

436

transition metals-mediated PMS process (in the absence of gamma-ray) was investigated

437

through radical scavengers study using alcohols. This study was, however, conducted

438

only for the most efficient Fe2+-mediated PMS process in the absence and presence of

439

alcohols. According to Anipsitakis et al. [27], for alcoholic radical scavengers, alcohols

440

with α-hydrogens react effectively with both SO4•− and •OH, therefore, in the present

441

study, iso-propyl alcohol (i-PrOH) was used as a scavenger for both SO4•− and •OH

442

(reactions (26) and (27) in Table 1) while 2-BuOH was used as a scavenger of •OH

443

(reaction (5) in Table 1). At a reaction time of 2 hours (equivalent to gamma-ray

444

absorbed dose of 592 Gy), the removal efficiency of chlorendic acid by PMS/Fe2+was

445

38% and 35% in the presence of 2-BuOH and i-PrOH, respectively, as compared to 70%

446

in the absence of alcohols (Figure 4). The lower removal efficiency of chlorendic acid by

447

PMS/Fe2+ in the presence of 2-BuOH and i-PrOH reveal yield of •OH and SO4•− from the

448

activation of PMS by transition metals and further confirm involvement of both •OH and

449

SO4•− in the removal of chlorendic acid.

450

3.8.

451

The decomposition of chlorendic acid by gamma-ray irradiation under the studied

452

conditions is expected to result in several transformation products (TPs) including larger

Formation of transformation products from chlorendic acid

20

453

molecules and short chain organic acids as well as inorganic ions. The decomposition of

454

chlorendic acid by SO4•− and •OH led to the formation of chlorendic anhydride, 1,3-

455

cyclopentadiene, 1,2,3,4,5,5-hexachloro, short chain organic acid, i.e., acetate ion, and

456

inorganic chloride ion as shown in Table 4 with their respective retention time (RT),

457

molar mass (MW), structural formula along with analytical techniques used for their

458

determination. All the TPs were found to result from the reaction of both SO4•− and •OH.

459

Both SO4•− and •OH react through common reaction mechanisms, i.e., hydrogen

460

abstraction and electron transfer, and therefore resulted in identical TPs in our study [9,

461

12]. The transfer of chlorendic acid to chlorendic anhydride by SO4•− and •OH possibly

462

take place through hydrogen abstraction reaction from carboxylate group involving the

463

loss of H2O molecule and yield of oxyl radical [28] which is then quickly converted into

464

chlorendic anhydride possibly through several intermediate steps following the loss of

465

H2O2 molecule as shown in Scheme 1(A).

.OH/SO .-

.OH/SO .-

4

4

-

H2O/HSO4-

466 467

Chlorendic acid

-H2O2/HSO5-

Chlorendic anhydride Oxyl radical

Scheme 1(A): Proposed degradation pathway for the formation of chlorendic anhydride

468

The decomposition of chlorendic anhydride by SO4•− and •OH at the position

469

identified by wedge-dash bond yielded 1,3-cyclopentadiene, 1,2,3,4,5,5-hexachloro with

470

the loss of furan-2,5-dione (Scheme 1(B)) [12].

471

21

472 Cl

Cl

O Cl

. OH/SO . -

Cl

Cl

Cl

4

O

-

Cl

Cl Cl

O

473

Cl

O

Cl

Cl

O

O

1,3-Cyclopentadiene, 1,2,3,4,5,5-hexachloro

Chlorendic anhydride

474

Scheme 1(B): Proposed pathway of 1,3-cyclopentadiene, 1,2,3,4,5,5-hexachloro

475

formation

476

The step wise reactions of 1,3-cyclopentadiene, 1,2,3,4,5,5-hexachloro with SO4•−

477

and •OH followed by hydrolysis led to ring opening and consequent yield of chloride

478

(Cl−) and acetate (CH3CO2−) ions as shown in Scheme 1(C) [9]. Cl

Cl

Cl

Cl

Cl

Cl Cl

.OH/SO .-

Cl

4

Cl

Cl

Cl

Cl Cl

Cl

H2 O

-Cl

.

- OH

Cl Cl

* HO

Cl

+

1,3-Cyclopentadiene, 1,2,3,4,5,5-hexachloro

OH 2

-H+

Cl

Cl.OH/SO .-

Cl Cl

4

Ring opening Cl Cl

C

3CO2

-

OH

479 480

Scheme 1(C): Proposed pathway for the formation of chloride and acetate ions

481

The formation of short chain organic acid, i.e., acetate ion reveals potential

482

decrease in mass concentration of the target contaminant [21]. Except 1,3-

483

cyclopentadiene, 1,2,3,4,5,5-hexachloro, the other three transformation products were

484

successfully quantified at the studied experimental conditions. The reactions of SO4•−

485

and •OH led to a steady decrease in concentration of chlorendic acid with the increase in 22

486

absorbed dose, however, concentrations of chlorendic anhydride and acetate initially

487

increased and then decreased continuously after extended treatment (Figure 5). The

488

decrease in concentrations of chlorendic anhydride and acetate ion after extended

489

treatment suggests contribution of reactive radicals in the removal of TPs as well.

490

Besides, this study verifies competition between the target contaminant and TPs for

491

reactive radicals.

492

The dechlorination of chlorendic acid, a hexachlorinated organochlorine

493

compound resulted in continuous loss of chloride ion with rapid increase in concentration

494

of chloride ion upto certain extent and then slowed down afterwards. The initial

495

concentration of 1.4 µM of chlorendic acid, a hexachlorinated organochlorine compound

496

is expected to yield 8.4 µM of chloride ions; however 4 µM chloride ions, corresponding

497

to a mass balance of 47.6% loss of chloride ions, was formed at an absorbed dose of 1000

498

Gy in the present study (Figure 6). When the absorbed dose was increased to 3000 Gy, 7

499

µM chloride ions, corresponding to a mass balance of 83% loss of chloride ion, was

500

formed (Figure 6). The rapid formation of chloride ion initially and leveling off afterward

501

suggest the formation of some persistent aliphatic compounds containing chlorine. .

502

The successful loss of chloride ion from chlorendic acid, responsible for toxicity

503

of chlorinated organic compounds, achieved in the present study suggest that significant

504

toxicity reduction of the water contaminated with chlorendic acid and related

505

organochlorine compounds could be achieved by gamma radiation treatment [9, 15].

506

23

507 508

4. Conclusions

509

Gamma radiation was successful in the removal of chlorendic acid and removal

510

efficiency was significantly improved in the presence of PMS. The removal efficiency of

511

chlorendic acid by gamma-ray/PMS was found to be due to the performance of •OH and

512

SO4•−. The yield of •OH and SO4•− from PMS in the presence of gamma radiation as well

513

as performance of •OH and SO4•− in the removal of chlorendic acid was investigated

514

through radical scavengers and competition kinetic study. The removal efficiency of

515

chlorendic acid increased with increasing initial PMS concentration and decreasing initial

516

target contaminant concentration. The removal efficiency of chlorendic acid was

517

inhibited in the presence of •OH and SO4•− scavengers, i.e., NO2− and CO32−, p-CBA, m-

518

TA, and alcohols. The presence of transition metals with gamma-ray/PMS process was

519

found to activate PMS and consequently increased the rate of formation of •OH and

520

SO4•−. The possible degradation pathways of chlorendic acid by •OH and SO4 •− was

521

proposed based on the degradation of chlorendic acid and nature of identified TPs. The

522

formation of acetate and chloride ions implicates potential decrease in mass concentration

523

of the target contaminant and detoxification of its aqueous solution. A decrease in mass

524

concentration of some of the TPs after extended treatment reveals significant competition

525

between parent compound and TPs for reactive radical. The efficient removal of

526

chlorendic acid suggests potential applications of gamma-ray with PMS process for

527

treatment of emerging organic contaminant in natural water.

528 529

24

530

Acknowledgment

531

The authors are thankful to the Higher Education Commission Pakistan (HEC) for

532

fellowship for higher study (to NSS) and research project grant (to HMK). The authors

533

are also thankful to the Nuclear Institute for Foods and Agriculture (NIFA) authorities for

534

permission to use gamma irradiation facility for this project.

535

25

536

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crosslinking and strand breakage of poly(viny1 alcohol) in aqueous solution in the

624

absence and presence of oxygen. A pulse radiolysis and product study, Macromol.

625

Chern. Phys. 195 (1994) 1443 -1461.

29

626

[29]

627 628

P. Maruthamuthu, P. Neta, Phosphate radicals, Spectra, acid-base equilibriums, and reactions with inorganic compounds, J. Phys. Chem. 82 (1978) 710-713.

[30]

G.V. Buxton, Pulse radiolysis of aqueous solutions. Some rates of reaction of OH

629

and O− and pH dependence of the yield of O3−, Trans. Faraday Soc. 65 (1969)

630

2150-2158.

631

[31]

632 633

P. Neta, R.E. Huie, A.B. Ross, Rate constants for reactions of inorganic radicals in aqueous solution, J. Phys. Chem. Ref. Data 17 (1988) 1027-1284.

[32]

G.G. Jayson, B.J. Parsons, A.J. Swallow, Some simple, highly reactive, inorganic

634

chlorine derivatives in aqueous solution. Their formation using pulses of radiation

635

and their role in the mechanism of the Fricke dosimeter, J. Chem. Soc. Faraday

636

Transac. 1 (1973) 1597-1607.

637

[33]

M.M. Abdel daiem, J. Rivera-Utrilla, R. Ocampo-Pérez, M. Sánchez-Polo, J.J.

638

López-Peñalver, Treatment of water contaminated with diphenolic acid by gamma

639

radiation in the presence of different compounds, Chem. Eng. J. 219 (2013) 371-

640

379.

641

[34]

W.J. McElroy, A laser photolysis study of the reaction of SO4•− with Cl− and the

642

subsequent decay of Cl2− in aqueous solution, J. Phys. Chem. 94 (1990) 2435-

643

2441.

644 645

[35]

V. Nagarajan, R.W. Fessenden, Flash photolysis of transient radicals. 1: X2− with X = Cl, Br, I, and SCN, J. Phys. Chem. 89 (1985) 2330-2335.

646

30

647 648

[36]

C.L. Clifton, R.E. Huie, Rate constants for hydrogen abstraction reactions of the sulfate radical, SO4•−. Alcohols, Int. J. Chem. Kinet. 21 (1989) 677-687.

649

31

650

Figure Captions

651

Figure 1. Effects of initial PMS concentration on the removal efficiency of chlorendic

652

acid by gamma-ray/PMS process. Experimental conditions: [chlorendic acid]0 = 1.40 µM,

653

pH = 5.2, gamma-ray dose rate = 296 Gy/h.

654 655

Figure 2. Effect of carbonate, sulfate, chloride, and nitrite ions on the removal efficiency

656

of chlorendic acid by gamma-ray/PMS process. Experimental conditions: [chlorendic

657

acid]0 = 1.40 µM, [PMS]0 = 140.0 µM, [CO32−]0 = [NO2−]0 = [Cl−]0 = [SO42−]0 = 1.0 mM,

658

pH = 5.2, gamma-ray dose rate = 296 Gy/h.

659 660

Figure 3. Removal of chlorendic acid by gamma radiation in the presence of PMS only,

661

and PMS with transition metals, i.e., cuprous, ferrous, and ferric ions. Experimental

662

conditions: [chlorendic acid]0 = 1.40 µM, [PMS]0 = 140.0 µM, [Cu+]0 = [Fe2+]0 = [Fe3+]0

663

= 14.0 µM, pH = 5.2, gamma-ray dose rate = 296 Gy/h.

664 665

Figure 4. Removal of chlorendic acid by PMS/Fe2+ in the absence and presence of i-

666

PrOH and 2-BuOH. Experimental conditions: [chlorendic acid]0 = 1.40 µM, [PMS]0 =

667

140.0 µM, [Fe2+]0 = 14.0 µM, [i-PrOH]0 = [2-BuOH]0 = 60 mM, pH = 5.2.

668 669

Figure 5. Changes in concentration of chlorendic acid, chlorendic anhydride, and acetate

670

with absorbed gamma-ray doses in the presence of PMS, Experimental conditions:

671

[chlorendic acid]0 = 1.40 µM, [PMS]0 = 140.0 µM, pH = 5.2, gamma-ray dose rate = 296

672

Gy/h.

32

673

Figure 6. Changes in the concentration of chlorendic acid and chloride ion with absorbed

674

doses in gamma-ray/PMS process, Experimental conditions: [chlorendic acid]0 = 1.40

675

µM, [PMS]0 = 140.0 µM, pH = 5.2, gamma-ray dose rate = 296 Gy/h.

676 677

Table Captions

678

Table 1. Rate constants for the potential elementary reactions in the gamma-ray based

679

AOPs under different conditions.

680 681

Table 2. Removal efficiency (%), observed pseudo-first-order degradation rate constants

682

(kobs, Gy−1), G-value (µmol/J), and dose required for 90% removal of chlorendic acid by

683

gamma–ray only (control), gamma-ray/PMS, gamma-ray/PMS/p-CBA, and gamma-

684

ray/PMS/m-TA processes. Experimental conditions: [chlorendic acid]0 = [p-CBA]0 = [m-

685

TA]0 = 1.40 µM, [PMS]0 = 140.0 µM, pH = 5.2 (in case of control pH was 5.8), gamma-

686

ray dose rate = 296 Gy/h.

687 688

Table 3. Effects of chlorendic acid initial concentration on its removal efficiency (%),

689

observed pseudo-first-order degradation rate constant (kobs, Gy−1), initial degradation rate

690

(µM/Gy), and G-value (µmol/J) in the presence of PMS. Experimental conditions:

691

[chlorendic acid]0 = 0.35, 0.70, 1.40, and 2.80 µM, [PMS]0 = 140.0 µM, pH = 5.2,

692

gamma-ray dose rate = 296 Gy/h.

693 694

Table 4. List of transformation products formed during the degradation of chlorendic acid

695

by gamma radiation with PMS based processes.

696 33

697

Figure 1

PMS = 35.0 µΜ PMS = 70.0 µΜ PMS = 140.0 µΜ PMS = 280.0 µΜ

1.0

C/C 0

0.8

0.6

0.4

0.2

0.0 0

698

200

400

600

800

1000

Absorbed dose (Gy)

34

699 700

Figure 2

PMS/NO2PMS/CO32PMS/Cl-

1.0

PMS/SO42-

C/C 0

0.8

PMS only

0.6

0.4

0.2

0.0 0

701

200

400

600

800

1000

Absorbed dose (Gy)

702

35

703

Figure 3

PMS only PMS/Fe3+ PMS/Cu+ PMS/Fe2+

1.0

C/C 0

0.8

0.6

0.4

0.2

0.0 0

704

200

400

600

800

1000

Absorbed dose (Gy)

705 706

36

707

Figure 4

2+

PMS/Fe /i-PrOH PMS/Fe2+/2-BuOH 2+ Fe /PMS

1.0

C/C 0

0.8

0.6

0.4

0.2

0.0 0.0

708

0.4

0.8

1.2

1.6

2.0

Time (h)

709

37

Figure 5

Chlorendic acid Chlorendic anhydride Acetate ion

1.4

0.8

Chlorendic acid (µΜ )

1.2 1.0

0.6

0.8 0.4

0.6 0.4

0.2 0.2 0.0

0.0 0

711

Chlorendic anhydride/Acetate ion (µΜ )

710

200

400

600

800

1000

Absorbed dose (Gy)

712

38

713

Figure 6

Chlorendic acid Chloride ion

1.4

7.0

5.6 1.0 4.2

0.8 0.6

2.8

Chloride ion (µΜ)

Chlorendic acid (µΜ)

1.2

0.4 1.4

0.2 0.0

0.0 0

714

500

1000

1500

2000

2500

3000

Absorbed dose (Gy)

39

715

Table 1

Rate constants (M−1s−1)

No. Reaction

References

1

eaq− + HSO5− → •OH + SO42−

8.4 × 109

[19, 20]

2

eaq− + HSO5− → SO4•− + −OH

8.4 × 109

[19, 20]

3

•

OH + p-chlorobenzoic acid → Product

5.0 × 109

[12]

4

SO4•− + m-Toluic acid → Product

2.0 × 109

[12]

5

•

3.1 × 109

[23]

6

•

H + 2-BuOH → CH3CH2•C(OH)CH3 + H2

1.2 × 108

[23]

7

SO4•− + Oxetane → H-abstraction

1.1 × 108

[23]

8

•

H + 2-chlorobenzoic acid → Product

6.2 × 108

[23]

9

•

OH + NO2− → NO2• + OH─

8.0 × 109

[29]

10

SO4•− + NO2− → NO2• + SO42−

8.8 × 108

[30]

11

•

OH + CO3─ → CO3•− + −OH

4.0 × 108

[23]

6

4.1 × 10

[29]

3.5 × 105

[31]

OH + 2-BuOH → Product

•−

─

•−

2−

12

SO4 + CO3 → CO3 + SO4

13

•

14

SO4•− + SO42− → S2O82− + eaq−

15

•

OH + Cl− → ClOH•−

4.3 × 109

[32, 33]

16

SO4•− + Cl− → Cl• + SO42−

6.6 × 108

[34]

17

ClOH•− + H+ → Cl• + H2O

8.8 × 108

[32]

18

Cl• + Cl− → Cl2 •−

8.0 × 109

[35]

19

Cl• + H2O → H+ + •OH + Cl−

2.0 × 105

[31]

20

HSO5− + Cl− → SO42− + HOCl

21

Fe2+ + HSO5− → Fe3+ + SO4•− + OH−

3.0 × l04

[11]

22

Fe2+ + HSO5− → Fe3+ + SO42− + •OH

3.0 × l04

[11]

3

OH + SO42− → SO4•− + −OH

+

−

2+

•−

[23]

[14, 23]

−

23

Cu + HSO5 → Cu + SO4 + OH

6.0 × l0

[26]

24

Cu+ + HSO5− → Cu2+ + SO42− + •OH

6.0 × l03

[26]

25

Fe3+ + HSO5− → Fe2+ + SO5•− + H+

26

SO4•− + i-PrOH → (CH3)2•COH + SO42− + H+

27

•

•

OH + i-PrOH → (CH3)2 COH + H2O

[11] 8.2 × 107

[36]

9

[23]

1.9 × 10

716 40

717

Table 2 Reaction conditions

% degradation

kobs (Gy−1)

G-value (µmol/J)

D0.90 (Gy)

Gamma-ray/PMS

95.0

2.90 × 10−3

3.80 × 10−3

794

Control

82.0

1.70 × 10−3

3.00 × 10−3

1350

Gamma-ray/PMS/p-CBA

45.0

6.20 × 10−4

1.20 × 10−3

3710

Gamma-ray/PMS/m-TA

40.0

5.40 × 10-4

1.00 × 10-3

4260

718 719

41

720

Table 3 Concentration % degradation

kobs (Gy−1)

(µM)

Degradation rate

G-value (µmol/J)

(µM/Gy)

0.35

96.0

5.30 × 10−3

6.00 × 10−4

1.50 × 10−3

0.70

87.0

3.53 × 10−3

1.30 × 10−3

2.30 × 10−3

1.40

78.0

2.94 × 10−3

2.00 × 10−3

3.80 × 10−3

2.80

61.0

1.90 × 10-3

3.40 × 10-3

5.40 × 10-3

721 722 723

42

724 Table 4 725 S# Compound

Structural formula

MW

RT (min)

Analytical techniques applied

1

Chlorendic acid

388.8

14.5

GC-ECD

2

Chlorendic anhydride

370.8

12.7

GC-ECD

3

1,3-cyclopentadiene,

273.0

11.3

GC-ECD

1,2,3,4,5,5-hexachloro

4

Acetate ion

CH3COO−

59.0

7.6

IC

5

Chloride ion

Cl−

35.5

10.5

IC

726 727

43

Graphical Abstract

728 729 730

Slower to faster removal efficiency Radical Scavengers

Co-60

gamma radiation PMS

. OH

.

SO4 -

Transition metal catalyst

Chlorendic acid, C H Cl O Chlorendic acid, C 99H44Cl66O44

ClCH3CO2-

731 732

44

Highlights

733 734 735 736


The presence of HSO5− with gamma-ray promoted removal efficiency of chlorendic acid.

737


The activation of HSO5− by gamma-ray and catalyst yield •OH and SO4•−.

738


The radical scavengers inhibited the efficiency of •OH and SO4•−.

739


Second-order rate constants of chlorendic acid with eaq−, •OH, and SO4•− were

740 741

determined.
Degradation pathways were proposed from the nature of identified by-products.

742 743

45