Formation and Stabilization of Environmentally Persistent Free

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Formation and stabilization of environmentally persistent free radicals induced by the interaction of anthracene with Fe(III)-modified clays Hanzhong Jia, Gulimire Nulaji, Hongwei Gao, Fu Wang, Yunqing Zhu, and Chuanyi Wang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b00527 • Publication Date (Web): 25 May 2016 Downloaded from http://pubs.acs.org on May 29, 2016

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Environmental Science & Technology

1

Formation and stabilization of environmentally persistent free

2

radicals induced by the interaction of anthracene with

3

Fe(III)-modified clays

4

a, b

Hanzhong Jiaa, Gulimire Nulaji

a

a

a

, Hongwei Gao , Fu Wang , Yunqing Zhu , and Chuanyi Wanga *

5 a

6

Laboratory of Environmental Sciences and Technology, Xinjiang Technical Institute of

7

Physics & Chemistry; Key Laboratory of Functional Materials and Devices for Special

8

Environments, Chinese Academy of Sciences, Urumqi 830011, China.

9

b

School of Geology and Mining Engineering, Xinjiang University, Urumqi 830046, China.

10 11 12 13 14 15 16 17 18 19

*To whom correspondence should be addressed.

20 21 22 23 24 25

Xinjiang Technical Institute of Physics & Chemistry Chinese Academy of Sciences 40-1 South Beijing road, Urumqi, Xinjiang, 830011, China Phone: +86-911-3835879 Fax: +86-911-3838957 E-mails: [email protected] (CYW); [email protected] (HZJ)

26 1

ACS Paragon Plus Environment


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ABSTRACT:

28

detected in Superfund sites but the formation of EPFRs induced by polycyclic

29

aromatic hydrocarbons (PAHs) is not well understood. In the present work, the

30

formation of EPFRs on anthracene-contaminated clay minerals was quantitatively

31

monitored

32

surface/interface-related environmental influential factors were systematically

33

explored. The obtained results suggest that EPFRs are more readily formed on

34

anthracene-contaminated Fe(III)-montmorillonite than in other tested systems.

35

Depending on the reaction condition, more than one type of organic radicals including

36

anthracene-based radical cations with g-factors of 2.0028-2.0030 and oxygenic

37

carbon-centered radicals featured by g-factors of 2.0032-2.0038 were identified. The

38

formed EPFRs are stabilized by their interaction with interlayer surfaces, and such

39

surface-bound EPFRs exhibit slow decay with 1/e-lifetime of 38.46 days.

40

Transformation pathway and possible mechanism are proposed on the basis of

41

experimental results and quantum mechanical simulations. Overall, the formation of

42

EPFRs involves single-electron-transfer from anthracene to Fe(III) initially, followed

43

by H2O addition on formed aromatic radical cation. Due to their potential exposure in

44

soil and atmosphere, such clay surface-associated EPFRs might induce more serious

45

toxicity than PAHs and exerts significant impacts on human health.

46

KEYWORD:

47

Environmentally persistent free radicals (EPFRs); Clay minerals; Polycyclic aromatic

48

hydrocarbons (PAHs); Electron transfer; Surface interaction.

via

Environmentally persistent free radicals (EPFRs) are occasionally

electron

paramagnetic

resonance

(EPR)

spectroscopy,

and

2


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49

TOC Art

50 51

INTRODUCTION

52

Environmentally persistent free radicals (EPFRs) are considered as a new class of

53

emerging pollutants due to their potential of inducing the formation of biologically

54

damaging reactive oxygen species (ROS), which may be responsible for the oxidative

55

stress causing cardiopulmonary disease and probably cancer.1 EPFRs have been

56

previously observed in combustion-generated particles and airborne particulate matter

57

with d < 2.5 µm (PM2.5).2 These EPFRs are produced by substituted aromatic

58

molecules (e.g., chlorophenols and chlorobenzenes) on the surfaces of transition

59

metal-containing particles at temperatures between 150 and 500 oC.2 The thermal

60

reaction processes are accomplished within a few seconds, but the formed EPFRs

61

persist in ambient air for days.3,

62

pentachlorophenol-contaminated soil from a former wood treatment facility sites.5

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The reactions are quite facile, occurring at room temperature, and have also been

64

found in Superfund sites contaminated with polycyclic aromatic hydrocarbons

65

(PAHs), polychlorinated biphenyls (PCBs), and polybrominated biphenyl ethers

66

(PBDEs).5-7 This inspired us to consider that EPFRs might be more ubiquitous than

67

previously suspected or envisioned, especially at sites contaminated with organic

68

pollutants.

4

Recently, EPFRs were detected in

3



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In soil phase, the formation of EPFRs correlates with the interaction between

70

selective aromatic compounds and soil components, such as inorganic minerals, soil

71

organic matter, and the biological components.8-10 Sequestering and binding of

72

organic contaminations to clay minerals play a significant role in their transformation

73

and persistence.11 Smectite, including montmorillonite, is a representative clay

74

mineral, which generally consists of a center octahedral Al-O sheet sandwiched

75

between two tetrahedral Si-O sheets. The planar aluminosilicate layers typically exist

76

in stacked assemblages, which are often referred to as tactoids. The unique properties,

77

such as negatively charged layers, high cation exchangeable capacity (CEC), and

78

expansible interlayer spaces, enable smectite to provide desired active sites for

79

organic pollutants bonding on its surface, thereby leading to various physicochemical

80

processes.12-14 Saturation of various cations is expected to modify the structural and

81

physicochemical properties of the clay minerals, and thus influences the interaction

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between organic pollutants and clay surfaces.15 When exchanged with certain

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transition metal ions (e.g., Cu(II) and Fe(III)), clay minerals can make a variety of

84

aromatic molecules, such as chlorinated phenols and anisoles, transform via interface

85

electron transfer, often followed by further reactions including dechlorination and

86

polymerization.6,

87

through the formation of organic radicals as intermediates with simultaneous

88

reduction of surface cations.7 Such organic radical intermediates are typically unstable

89

and short-lived species. However, surface-bound organic radicals associated with

90

mineral particles are occasionally persistent and are relatively long-lived in the

7, 16, 17

Transformation of these pollutants is possibly achieved

4


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environment, i.e., EPFRs.4

92

Besides substituted phenols or benzenes, PAHs could also be transformed on clay

93

surfaces.18 PAH molecules, which possess highly delocalized π-electrons, may act as

94

strong electron-donors when interacting with electron-deficient species such as

95

exchangeable cations via electron-donor–acceptor interactions.19-22 Such “cation-π”

96

interaction has been demonstrated as an important factor regulating PAHs availability

97

and transformation on mineral surfaces.12 Among commonly found cations (Na(I),

98

K(I), Ca(II), Mg(II), Al(III), and Fe(III)), the presence of transition metal ions such as

99

Fe(III) on clay surfaces facilitates the transformation of PAHs due to their strong

100

cation-π interactions.22 The transformation is accompanied by the electron transfer

101

from the aromatic species to surface cations.23-25 Therefore, free radicals are very

102

likely to be generated during the interaction between modified clay minerals and

103

molecular PAHs. The formed free organic radicals associated with the clay surfaces

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retain their additional stabilization, might allowing them to persist in the

105

environment.26, 27 However, limited work has been conducted to assess the formation

106

of PAHs-induced free radicals and their persistence on clay surfaces, and thus critical

107

information is missing for the evaluation of potential risks from PAHs-contaminated

108

soil in association with EPFRs.

109

In this work, we demonstrate for the first time the potential of EPFRs formation

110

induced by PAHs, and in particular, by anthracene on clay mineral surfaces under

111

environmentally relevant conditions. Conversion of molecular PAHs to EPFRs as

112

well as persistency of the EPFRs was monitored via electron paramagnetic resonance 5



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(EPR) analysis. The principal objectives of the work are to 1) probe the role of

114

interactions between PAHs and clay minerals in PAHs transformation and EPFRs

115

formation; 2) reveal the influence of ionization potential of the organic molecule and

116

clay surface properties on clay-mediated EPFRs formation; and 3) further gain insight

117

into the mechanisms of PAHs-induced EPFRs formation on Fe(III)-clay surface. This

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work will provide useful information for the evaluation of potential risks from

119

PAHs-contaminated clay minerals in association with free radicals.

120

EXPERIMENTAL SECTION

121

Chemicals and materials. Detailed information on the chemicals used in this

122

study is supplied in the Supporting Information (SI). Reference montmorillonite, illite,

123

and kaolinite were obtained from Zhejiang Feng-Hong Clay Chemicals Co., Ltd

124

(ZheJiang, China). These clays were dissimilar in CEC, specific surface area, and

125

interlayer swelling properties (Table S1).

126

EPFRs formation. PAHs-contaminated clay minerals were prepared by adopting a

127

protocol as previously reported.28 Briefly, clay minerals (< 2 µm) were obtained by

128

centrifugation of clay suspension for 6 min at 600 rpm, and then treated with 0.1 M

129

FeCl3 solution for four times. The clay samples before and after Fe(III) saturation

130

were digested by the mixture of hydrofluoric acid, nitric acid and perchloric acid at

131

250 oC for 90 min, and Fe contents were determined using a Perkin-Elmer PinAAcle

132

900T Atomic Absorption Spectrophotometer (Nor-walk, CT). The reaction mixtures

133

of 0.1 mg/g PAHs-contaminated clay minerals were prepared by mixing 1 g of

134

Fe(III)-modified clays with 5 mL of various PAHs in methanol solution, in which 6


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anthracene, phenanthrene, pyrene, and benzo[a]pyrene were, respectively, employed

136

in each sample. Methanol was used as solvent of PAHs to allow its evaporation under

137

ambient conditions.

138

One gram of obtained PAHs-contaminated clay was laid onto a Petri dish, and then

139

placed inside a desiccator without light irradiation to prevent any light-induced

140

chemical reactions. The relative humidity (RH) was controlled by the saturated salt

141

solutions in reaction cells. To conduct the anoxic reaction, the reaction system was

142

transferred into an anoxic chamber without free O2 and H2O molecules. To investigate

143

the effect of reaction temperature, the desiccators with reaction systems were

144

transferred into an oven controlled to various temperatures. At pre-selected intervals

145

(such as 1 d, 2 d, 3 d, 5 d, 8 d, and max. 35 d), the samples were sacrificed and

146

transferred into 50 mL Teflon centrifuge tubes. The residual PAHs and produced

147

EPFRs were extracted with 10 mL of extraction solution (mixture of 5 mL acetone

148

and 5 mL dichloromethane) and analyzed immediately.

149

EPR Characterization. All EPR measurements were performed using a Bruker

150

E500 EPR spectrometer at room temperature. Instrument and operating parameters

151

are as follows: center field, 3470 G; microwave frequency, 9.7 GHz; microwave

152

power, 2.0 mW; modulation frequency, 4.0 G; modulation amplitude, 4.0 G; sweep

153

width, 200 G; receiver gain, 3.54 * 104; time constant, 41.0 ms; sweep time, 167.7 s;

154

and three scans. Radical concentrations were calculated by comparing the signal peak

155

area, as derived from (∆Hp-p)2 multiplied by the relative intensity, to a

156

2,2-diphenyl-1-picrylhydrazyl standard. 7



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Products Analysis. PAHs were quantified using a Thermo Fisher Ultra 3000

158

HPLC equipped with a 25 cm × 4.6 mm Cosmosil C18 column. A 85:15 (v/v) mixture

159

of methanol:water was employed as mobile effluent. The flow rate was 1.0 mL min−1,

160

and the ultraviolet detector was set at 254 nm. The PAHs intermediate products were

161

identified using a Agilent 7890A-5975C gas chromatograph incorporated with a mass

162

spectrometer operated on a full scan mode (30-500 amu), where a HP-5MS capillary

163

column (length = 30 m; internal diameter = 250 µm; film thickness = 0.25 µm) was

164

employed. Helium was used as carrier gas at a flow rate of 1.2 mL/min with splitless

165

injection at 230 oC. The oven temperature was programmed from 80 oC to 200 oC (20

166

o

167

in the reacted system was measured with following procedures. The reacted

168

Fe(III)-clay sample was mixed with d.i. water using Vortex for 30 s. Then 0.5 mL of

169

suspension was collected and added to 1 mL of ferrozine solution (100 mM), and the

170

volume of the mixture was adjusted to 15 mL. The suspension was agitated for 2 h

171

and filtered through a 0.45 µm filter. Concentration of ferrozine-complexed Fe(II) was

172

measured by UV–Vis spectrophotometer at 562 nm.

C min−1, 2 min hold), and then to 260 oC (20 oC min−1, 2 min hold). Content of Fe(II)

173

Decay study. Kinetic studies were performed to determine the persistency of free

174

radicals in air. The samples were exposed to ambient air, and EPR signal was

175

measured periodically to determine the radical concentration as a function of time.

176

For reaction rate calculations, a first order kinetic expression was used:

177 178

−dR / dt = k[R] where R is the concentration of detected EPFRs. The 1/e lifetime ( t1/e ) of EPFRs for 8


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the first-order decay was evaluated as following:

180

Ln(R / R0 ) = −kt , and t1/e = 1/ k

181

Rate constant k was derived from the slop of the correlation between logarithm of

182

radical concentration change (R/R0) vs time, and 1/e lifetime was thereby derived.

183

Theoretical modeling. Density functional theory (DFT) calculations were carried

184

out to model the reaction energies and activation energies associated with the

185

proposed reaction pathway using Dmol3 program29, 30 from the Material Studio (MS)

186

of Dassault Systèmes Biovia Corp. The geometry optimizations of all intermediate

187

and transition state structures were performed using the Becke exchange31 plus

188

Perdew-Wang approximation functional32 within Generalized gradient approximation

189

(GGA). The energies were unscaled and zero-point corrected. Transition states were

190

located by performing relaxed potential energy surface scans followed by

191

implementation of a complete linear synchronous transit (LST) and quadratic

192

synchronous transit (QST) method.

193

RESULTS AND DISCUSSION

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EPFRs formation on Fe(III)-montmorillonite. The potential generation of EPFRs

195

on Fe(III)-montmorillonite and Na(I)-montmorillonite contaminated by various PAHs

196

was studied by EPR under relatively dehydrated condition (RH ~8%). The

197

non-polluted clay minerals and Na(I)-montmorillonite contaminated by various PAHs

198

do not show any detectable EPR signals (Fig. S1). For Fe(III)-montmorillonite

199

systems, significant EPR signals were observed in anthracene-contaminated

200

Fe(III)-montmorillonite (Figs. 1a and S1). The total EPFRs yields increase with 9



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reaction time in 10 d, and then gradually decrease with reaction time (Fig. 1b).

202

Interestingly, more than 30% of EPR signals remain even after one month, indicating

203

the persistent nature of these detected free radicals. However, the EPFRs were not

204

observed in the system of Fe(III)-montmorillonite contaminated by other PAHs such

205

as phenanthrene, pyrene, and benzo[a]pyrene (Fig. S1). Formation of EPFRs might

206

relate to the PAHs transformation on clay surface. As shown in Fig. S2a, almost 65 %,

207

90 %, and 100 % of anthracene, pyrene, and benzo[a]pyrene are transformed in 3 d by

208

Fe(III)-montmorillonite, while insignificant changes in phenanthrene are observed

209

during the experimental period. For control experiments performed with

210

Na-montmorillonite, changes in PAHs were undetectable, implying that adsorption

211

and microbiological reactions therein were negligible in the present work time frame.

212

The varied transformation rates for individual PAHs can be ascribed to the fact that

213

PAHs with ionization potential lower than 7.6, such as anthracene (7.44 eV), pyrene

214

(7.43 eV), and benzo[a]pyrene (7.12 eV), readily undergo a single-electron transfer

215

reaction between PAHs and Fe(III) on clay surfaces18, 33 This is further supported by

216

the

217

PAHs-Fe(III)-montmorillonite system (Fig. S2b). In this electron-transfer process,

218

organic radical intermediates might be produced. However, the EPFRs cannot be

219

detected in the reaction systems associated with pyrene and benzo[a]byrene, which

220

are perhaps due to their less persistency making them hardly observable except at

221

very short time. On the other hand, Fe(III)-montmorillonite is unable to degrade

222

PAHs, such as phenanthrene, with ionization potential > 7.6 (see Fig S2a).18 Thus,

increase

in

Fe(II)

concentration

with

reaction

time

in

10


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electron transfer appears to be unfeasible in phenanthrene-Fe(III)-montmorillonite

224

system. Therefore, among those four PAHs, persistent free radicals can only be

225

observed on Fe(III)-montmorillonite contaminated by anthracene.

226

The g-factor of EPR signal is a useful parameter for identifying the type of free

227

radicals.34, 35 As shown in Fig. 1b, the g-factor in a range from 2.0033 to 2.0036

228

increases with reaction time initially and decreases afterwards. In addition, the

229

asymmetrical EPR spectral profiles indicate more than one type of EPFRs generated.

230

The de-convolution of the EPR spectra implies 3 different spectral components

231

therein, denoted as g1, g2, and g3 with the g-factors of 2.0028-2.0030, 2.0036-2.0038,

232

and ~ 2.0032, respectively (Fig. S3). Previous studies suggest that the PAHs-based

233

radical cations with g-factor of ~ 2.0028 are readily formed through the direct

234

electron-transfer from aromatic compounds to transition-metal ions on clay surface.36,

235

37

236

cations. The oxygen-centered radicals, such as semiquinone radical anions, possess a

237

g-factor > 2.0040. 35 Radical signals with g-factors of 2.0030-2.0040 are attributed to

238

oxygenic carbon-centered EPFRs or/and carbon-centered radicals with a nearby

239

heteroatom, such as oxygen or halogen, which increase the spin-orbit coupling

240

constant.38, 39 In the present study, therefore, the produced g2 and g3 signals were

241

originated from carbon-centered EPFRs with an adjacent oxygen atom or/and

242

oxygenic carbon-centered EPFRs. As shown in Figs. 1c and S3, the EPR spectra

243

shape and peak areas of the g1, g2, and g3 signals vary with reaction time. Initially,

244

the free radical of g1 dominates the EPR spectra within 2 d. After that, the peak area

Thus, the g1 EPR signal can be ascribed to the formation of anthracene-type radical

11



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of g1 signal gradually decreases to undetectable level in 6 d, while the peak area of g2

246

signal increases rapidly in the same time frame. The result suggests that the formation

247

of g2 radical can be ascribed to the decomposition of aromatic radical cations, which

248

is in agreement with previous studies.40 After 10 d, the peak area of g2 signal

249

generally decreases with further increase in reaction time, accompanied by the

250

increase in the yield of g3 signal. This suggests that the in situ-formed EPFRs can be

251

a precursor to form a new type of carbon-centered radicals with g-factor of ~ 2.0032

252

(Fig. 1c). After 17 d, most of g2 radicals are consumed. Meanwhile, the g3 signal is

253

also gradually decreased with increasing reaction time, implying the decomposition of

254

g3 radicals to the final product. The development of free radicals correlates with the

255

transformation processes of anthracene. The analyses of the GC-MS extracts of

256

anthracene-contaminated Fe(III)-montmorillonite are presented in Fig. S4. The main

257

transformation products of anthracene could be identified as anthraquinone formed by

258

ketonizing the intermediate benzene ring of anthracene (Fig. S5), while anthrone is

259

considered as a major detectable intermediate for the transformation from anthracene

260

to anthraquinone, suggesting the possible formation of carbon-centered EPFRs with

261

an adjacent oxygen atom (such as g2 radical) by partially ketonizing the benzene ring

262

of anthracene. Thus, the transformation from g1 radical to g2 radical is probably

263

accompanied with the formation of anthrone from anthracene. On the other hand, g3

264

radical might be involved in the ketonizing of the other carbon (i.e., C10) in the

265

intermediate benzene ring of anthracene, which may result in the formation of final

266

product, i.e., anthraquinone. 12


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In addition, the obtained three types of EPFRs exhibit varied persistency on the clay

268

surfaces. The decays of g1, g2, and g3 signals are depicted from the reaction time at

269

their highest yield, i.e., 3 d for g1, 10 d for g2, and 15 d for g3. As shown in Fig. 1d,

270

g1 signal exhibits a “fast” decay with 1/e lifetime of 1.41 d, indicating that aromatic

271

radical cations are relatively unstable on Fe(III)-saturated clay surfaces. The half-lives

272

of carbon-centered EPFRs with an adjacent oxygen atom, such as g2 radical, are

273

much longer than the corresponding radical cation species on Fe(III)-montmorillonite,

274

with a 1/e lifetime of 3.45 days. On the other hand, the g3 radical exhibits the

275

“slowest” decay, with a 1/e lifetime of 38.46 d (Fig. 1d). As reported previously, the

276

carbon-centered radical may remain stable when the para positions of the benzene

277

ring are occupied by stable substituents.17 Thus, the g3 signal with “slower” decay

278

and lower g-factor than g2 signal might be attributed to the formation of

279

carbon-centered radicals with para oxygen atom, such as anthroxyl radical.

280

Role of clay minerals on EPFRs formation. Although common free radicals are

281

typically unstable and of short-lived species, some free organic radicals generated at

282

or near the solid particle may have strong interactions with the particle surfaces,

283

thereby making them stable and persistent.41,

284

minerals on EPFRs formation, various Fe(III)-saturating clay minerals (i.e.,

285

Fe(III)-kaolinite,

286

anthracene was studied by EPR. As shown in Table S1, the CEC of montmorillonite is

287

greater than that of other clays, and the exchangeable cations are located on both

288

external and interlayer surfaces of montmorillonite. Kaolinite clay has essentially no

Fe(III)-illite,

and

42

To investigate the role of clay

Fe(III)-montmorillonite)

contaminated

by

13



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isomorphic substitution and the small amount of negative charges result mostly from

290

the dissociation of hydroxides at edge sites, hence a small quantity of surface Fe(III)

291

resides primarily on the edge sites. Though negatively-charged illite has high surface

292

density, most of structural charges in clay interlayers are compensated by fixed K+

293

which cannot be replaced by the added Fe(III), thus surface Fe(III) ions are mainly

294

located on external surfaces. Therefore, EPFRs formation on Fe(III)-saturated

295

kaolinite was used to examine the reactive sites on edge surfaces, while illite was used

296

to reflect the reactivity on external surfaces. Compared with kaolinite and illite, Fe(III)

297

located on montmorillonite could refer to the reactive sites in clay interlayers.

298

As displayed in Table S1, the surface Fe(III) contents were ca. 3.12 %, 1.54 %, and

299

0.44 % for Fe(III)-saturated montmorillonite, illite, and kaolinite, respectively, which

300

is derived by the difference in Fe content between original clays and Fe(III)-saturated

301

clays. However, the amount of EPFRs formed on anthracene-contaminated

302

Fe(III)-montmorillonite is > 4 orders of magnitude greater than that on Fe(III)-illite

303

and Fe(III)-kaolinite clays (Fig. S6). In other words, the difference in surface Fe(III)

304

content for the three Fe(III)-saturated clay systems is less than 1 order of magnitude,

305

while the difference in EPFRs yields is more than 4 orders of magnitude. Therefore,

306

microenvironment of Fe(III) located in the clay interlayers plays vital role in EPFRs

307

stabilization among the reactive Fe(III) sites on clay surfaces, which agrees well with

308

what reported previously.16, 43 Generally, the EPFRs are located at specific sites on

309

mineral surfaces, in which stable free radicals are readily formed or in which the

310

produced radicals are easily stabilized.37 During the PAHs transformation process, the 14


Page 15 of 32


311

electron-transfer reaction could induce the formation of organic radical cations and

312

their oxygenic radicals.44, 45 The formed positively charged species would be strongly

313

bound to the negatively charged silicate surface.37, 46, 47 The large electric fields in the

314

interlayer region of clays favor an optimum dispersal of surface charge, which lowers

315

the electrostatic energy of various interaction or complexes in layered silicates.37 Thus,

316

free radicals formed from selected organic molecules at localized sites on clay

317

surfaces are more stable in clay interlayer than in outer spaces. On the other hand, the

318

aromatic molecules are tend to orientate in the interlayer region, which are favorable

319

in electron-transfer reactions and free radicals formation.37 Therefore free radicals

320

located at interlayer sites are relatively long-lived in the environment, i.e.,

321

environmentally persistent.

322

Effect of environmental condition

323

EPFRs formed in anoxic condition. The formation of EPFRs may be affected by

324

atmospheric O2 or/and H2O molecules, which may participate in the oxidation of

325

organic contaminants or/and the decomposition of free radical intermediates.4 To

326

explore the role of O2/H2O on EPFRs formation, anthracene transformation on

327

Fe(III)-montmorillonite clay was initially conducted in an anoxic chamber without

328

free O2 and H2O molecules. Under anoxic condition, the yield of EPFRs gradually

329

increases up to 15 d (Fig. 2). The de-convolution of their EPR spectra indicates that

330

the g1 signal, which is defined as aromatic radical cations, is initially formed and

331

relatively stable under anoxic condition compared to that under ambient condition at

332

RH of ~ 8% (Fig. S7). With increasing reaction time, the produced radical cations are 15



Page 16 of 32

333

partially transformed to more oxygenic radicals such as carbon-centered radicals with

334

an adjacent oxygen atom, which is consistent with observed increasing in g-factor

335

from 2.0028 to 2.0032 (Figs. 2 and S7). Although O2 and H2O molecules are limited

336

in an anoxic chamber, Fe(III) on clay surfaces may combine with OH- and/or H2O

337

molecules and readily form small oligomers such as [Fe(OH)1-4]n-1~2+ during its

338

preparation. As reported previously, the hydroxyl group from available H2O

339

molecules is incorporated into the radical products, and the reaction of radical cations

340

with OH- is dependent on the structure of precursor molecules, such as PAHs.33, 37

341

Radical cations localized at 9, 10 positions in the anthracene and similar compounds

342

react preferentially with OH- to result in quinines or/and diphenols as primary

343

products.33 Thus, the partial formation of oxygenic free radicals might be due to the

344

reaction between aromatic radical cations and OH-/water complexing with Fe(III) on

345

clay surfaces. Exposure of the radical cations, formed in anoxic chamber, to ambient

346

air (RH is ~ 8%) affects both their EPR signal positions and intensities. The g-factor

347

rapidly increases from 2.0032 to 2.0038 after prolonged air exposure (Fig. 2). Such

348

g-factor change indicates the trend of aromatic radical cations rapidly converting to

349

carbon-centered EPFRs with an adjacent oxygen atom.48 In addition, the peak area

350

shows little change for the first few days of air exposure, and then gradually decreases

351

after prolonged air exposure, accompanied with the decrease in g-factor. The result

352

indicates the gradual decomposition of carbon-centered radical with nearby oxygen to

353

the other carbon-centered radical and further generation of final product, i.e.,

354

anthraquinone. 16


Page 17 of 32


355

The effect of temperature. The transformation of organic pollutants on clay

356

surfaces and formation of EPFRs is likely to be temperature-dependent.49 Fig. 3a

357

depicts the derivative EPR spectra of EPFRs observed for 15 d transformation of

358

anthracene as a function of reaction temperature ranging from 25 to 90 °C. The total

359

yields of EPFRs exhibit insignificant changes with increasing reaction temperature

360

from 25 °C to 40 °C. However, further increase in reaction temperature induces

361

decreasing in radical yields (Fig. 3b). When raising temperature above 75 °C, the

362

EPFRs yields become approximately constant and relatively low, suggesting very

363

limited conversion of anthracene to its EPFRs. The increase in the reaction

364

temperature reduces the yields of EPFRs, which might be due to either lower initial

365

yield of free radicals or high reactivity for the decomposition of free radicals on clay

366

minerals. It is noted that the transformation rate of anthracene is significantly

367

enhanced when the reaction temperature increases from 40 °C to 90 °C (Fig. 3c). The

368

obtained results indicate that the low yield of EPFRs can be mainly attributed to their

369

easier decomposition under higher temperature, inducing rapid conversion of

370

anthracene to its final product.

371

Although a slight change of EPR g-factors was observed under various reaction

372

temperatures, interestingly, the g-factor gradually increases from 2.00335 to 2.00351

373

as the reaction temperature increases from ~ 40 to 90 °C (Fig. 3b). The

374

temperature-dependent evolution of g-factor is related to the type and relative yields

375

of EPFRs. As discussed above, newly formed oxygenic carbon-centered EPFRs (i.e.,

376

g3 radical) dominates the EPR spectra at 25 oC, accompanied by the formation of a 17



Page 18 of 32

377

small amount of g2 radical (i.e., carbon-centered EPFRs with an adjacent oxygen

378

atom). The result suggests a higher contribution of g3 radical at ≤ 40 °C, while

379

oxidation or losing of more g3 radicals readily occurs at higher temperature. Thus,

380

carbon-centered radical with adjacent oxygen atom might be the predominant species

381

at relatively high reaction temperature, which induce the increase in g-factor of

382

EPFRs. This is also consistent with the change of EPFRs yields at various reaction

383

temperatures, in which less EPFRs yields were observed at > 40 °C (Fig. 3b).

384

Meanwhile, increase in the EPFRs yields at ambient temperatures such as 25-40 °C

385

simply indicates that the transformation rate of anthracene exceeds the rate of

386

carbon-centered radical decomposition on clay surface.

387

The effect of relative humidity. Besides the reaction temperature, the nature and

388

amount of produced free organic radicals also depends on the RH in the reaction

389

medium.37 As shown in Figs. 3d and S8, the increase in the RH that ranges from 8%

390

to 11% leads to a small amount of improvement in EPFRs yields accompanied with

391

increased transformation rate of anthracene. This might be due to that the ligand water

392

molecules around surface cations participate in the EPFRs formation reaction.33, 37

393

Further increase in RH above 11% results in a steep decrease in both of the EPFRs

394

yields and PAHs transformation rate. When RH is > 43%, the transformation of

395

anthracene to EPFRs on Fe(III)-montmorillonite surfaces is almost completely

396

retarded (Fig. S8). This result is in agreement with previous report, in which the

397

addition of water to a system associated with arene transformation on silica-alumina

398

resulted in a rapid decay of reaction rate.50 The suppress effect by water is attributed 18


Page 19 of 32


399

to a competition between arene and water molecule for the Lewis acid site (such as

400

Fe(III) on clay surfaces).50 Generally, cations on clay surfaces tend to be hydrated,

401

resulting in the formation of water layer in interfacial region.14 The coverage of water

402

molecules leads to an increased detachment of anthracene from the inner-sphere

403

coordination sites of cations, which influences the interaction between organic

404

contaminants and clay surfaces.22 The decreased anthracene-Fe(III) interaction on

405

clay surface induces the decrease in electron transfer reaction rate and EPFRs

406

formation

407

Fe(III)-montmorillonite enhances PAHs chemisorptions, which, in turn, allows the

408

reaction of electron transfer.51 Therefore, the presence of free water blocks the active

409

sites and hinders the catalytic effect of the clay surface; the interlayer water must be

410

removed to certain extent for the oxidation reaction of organic contaminants and the

411

formation of EPFRs to proceed.

on

clay

surface.44

On

the

other

hand,

dehydration

of

412

Theoretical prediction of possible EPFRs formation mechanism. Reaction of

413

aromatic compounds with montmorillonite saturated by transition metal cations (e.g.,

414

Fe3+ and Cu2+) has been previously studied.6, 52, 53 The results lead to the proposal that

415

during the reaction, electrons were donated by the unsaturated organic compounds to

416

the surface cations, resulting in the formation of aromatic radical cations and reduced

417

metal ions such as Fe(II) and Cu(I).6, 7 Similar to the mechanism proposed in previous

418

studies, PAHs transformation and EPFRs formation are also due to the electron

419

transfer process between PAHs and surface cations.18, 45 In order to further understand

420

the mechanism of EPFRs formation, the initial Fe(III) impact on the anthracene 19



421

Page 20 of 32

process is modeled by the transformation state theory.

422

Potential energy surface (PES) and optimized geometries of the anthracene reaction

423

with Fe(III) are shown in Fig. 4. Fe(III) and anthracene form a pre-reaction complex,

424

IM1, by means of two intermolecular bonding (4.282 Å) interaction. The Fe(III)

425

addition to the C9 or C10 atom on the anthracene molecule processes through the

426

transition state TS with an activation barrier of 0.355 kcal/mol. This reaction can be

427

considered as a barrierless pathway. As a result of stabilizing the resonance ring

428

structure, the adducts of Fe(III) and anthracene leads to the formation of the

429

intermediate IM2 with 22.591 kcal/mol lower energy than the reactants, thereby

430

resulting in the formation of cation–π complexes on clay surfaces. In this process,

431

attack by Fe(III) at the C9 or C10 atom leads to –Fe···C– distance of 2.12 Å. Then,

432

the initial single-electron-transfer reaction is exoenergetic by 0.142 kcal/mol. The

433

associated electron-transfer within the complex results in the formation of

434

anthracene-type radical cation (radical A) and reduction of transition metal ions (as

435

shown in Scheme 1).36 The redox reaction is mainly facilitated and enhanced by the

436

planar negatively charged silicate layers of the montmorillonite clay.36 Although the

437

formed radical cations are stabilized on clay interlayer surface, the unpaired electrons

438

in the free radicals can be oxidized by atmospheric oxygen or hydrolyzed by H2O

439

molecules, resulting in the formation of other intermediate products.7, 16 According to

440

theoretical simulation, O2 and OH- can be added to C9 (or C10) atom of

441

anthracene-type radical cation via the transition state TS with an activation barrier of

442

1.803 and 0.118 kcal/mol, respectively. Finally, the –O···C– distance decreases from 20


Page 21 of 32


443

1.869 to 1.544 in the case of O2 addition and from 2.76 to 1.46 Å for OH- addition,

444

inducing the formation of final oxygenic radicals with energy of -15.618 and -30.777

445

kcal/mol, respectively. Thus, the reaction between H2O and radical A is more readily

446

than that between O2 and radical A, resulting in the formation of radical B (Scheme 1).

447

As

448

Fe(III)-montmorillonite, the radical B (hydroxyanthracene-type radical) deprotonates

449

to form the 9-hydroxy-anthracene, which is further rapidly tautomerized to the

450

thermodynamically favored anthrone (Scheme 1).54 Then, electron transfer between

451

anthrone and Fe(III)-montmorillonite induces the formation of anthrone-type radical

452

(radical C), followed by hydrolysis to generate oxanthrone, which can be further

453

transformed to anthraquinone.55 Radical B might also be directly transformed to

454

radical C through the intramolecular electron-transfer and deprotonation processes

455

accompanied by the formation of the intermediate products A and B. The present

456

study indicates that radical A is relatively unstable in natural environment and readily

457

transform to radical B. Compared with radical B, radical C (a carbon-centered

458

anthroxyl radical) are more stable due to the para positions of the benzene ring

459

occupied by ketone group.17

460

Environmental importance

the

pathway

proposed

for

anthracene

oxidative

degradation

on

461

The precursor molecules (PAHs) employed in this study are produced worldwide

462

via fossil fuel deposit and incomplete combustion, have been recognized as one class

463

of primary contaminants in naturally hydrophobic phases such as soil and sediment.56

464

Previous work reports the detection of EPFRs in PAHs-contaminated soil under

21



Page 22 of 32

465

ambient environmental conditions.57 But limited evidence is available to prove its

466

formation process. This study provides, for the first time, experimental evidence for

467

EPFRs formation on the clay minerals contaminated by the PAHs such as anthracene

468

under environmentally relevant conditions, i.e., presence of water and ambient

469

temperature. This study implies that many PAHs-contaminated soils may be at risk

470

for EPFRs production. Moreover, large production and propensity for volatilization

471

could lead to ubiquitous atmospheric contamination by PAHs. This suggests that the

472

highly stable EPFRs on clay mineral particles might be transported in the atmosphere

473

for a long distance from the source, eventually participate in atmospheric reactions, or

474

directly exert health and environmental impacts.58

475

Previous studies suggest that clay-based system such as Fe(III)- and Cu(II)-smectite

476

may be useful in the alteration and degradation of aromatic molecules present in

477

waste or contaminated sites.59 Such clay-catalyzed electron-transfer reactions were

478

considered as detoxification of organic toxicants under mild reaction conditions.

479

However, the toxicity of EPFRs has not been measured during the transformation of

480

those organic contaminants. Biochemical and biomedical studies on Fe-EPFRs

481

complexes indicate that such surface-associated EPFRs can induce the formation of

482

biologically damaging ROS such as peroxide, superoxide, and hydroxyl radical by

483

redox-cycles process, which may be responsible for the oxidative stress causing

484

cardiopulmonary disease and probably cancer that has been attributed to the humane

485

exposure to clay particles.60 Thus, transformation of PAHs on clay minerals could

486

potentially give rise to more toxic PAHs-type EPFRs. Therefore, the potential 22


Page 23 of 32


487

environmental risks from PAHs-contaminated clay minerals should be re-evaluated

488

due to their association with EPFRs.

489

ASSOCIATED CONTENT

490

Supporting Information Available

491 492

Additional details as noted in text. This information is available free of charge via the Internet at http://pubs.acs.org.

493

AUTHOR INFORMATION

494

Corresponding Author

495

*Phone: +86-911-3835879; e-mails: [email protected].

496

Notes

497

The anthors declare no competing financial interest.

498

ACKNOWLEDGMENTS

499

Financial support by the National Natural Science Foundation of China (Grants No.

500

41571446 and 41301543), the West Light Foundation of Chinese Academy of

501

Sciences (2015-XBQN-A-03), the Xinjiang Program of Introducing High-Level

502

Talents (Y539031601), and the CAS Youth Innovation Promotion Association

503

(2016380) are gratefully acknowledged.

504

REFERENCES

505 506 507 508 509 510 511 512 513 514 515 516

(1) Khachatryan, L.; Vejerano, E.; Lomnicki, S.; Dellinger, B. Environmentally persistent free radicals (EPFRs). 1. generation of reactive oxygen species in aqueous solutions. Environ. Sci. Technol. 2011, 45 (19), 8559-8566. (2) Gehling, W.; Dellinger, B. Environmentally persistent free radicals and their lifetimes in PM2.5. Environ. Sci. Technol. 2013, 47 (15), 8172-8178. (3) Cormier, S. A.; Lomnicki, S.; Backes, W.; Dellinger, B. Origin and health impacts of emissions of toxic by-products and fine particles from combustion and thermal treatment of hazardous wastes and materials. Environ. Health. persp. 2006, 114 (6), 810-817. (4) Lomnicki, S.; Truong, H.; Vejerano, E.; Dellinger, B. Copper oxide-based model of persistent free radical formation on combustion-derived particulate matter. Environ. Sci. Technol. 2008, 42 (13), 4982-4988. (5) dela Cruz, A. L. N.; Gehling, W.; Lomnicki, S.; Cook, R.; Dellinger, B., Detection of 23



517 518 519 520 521 522 523 524 525 526 527 528 529 530 531 532 533 534 535 536 537 538 539 540 541 542 543 544 545 546 547 548 549 550 551 552 553 554 555 556 557 558 559 560

Page 24 of 32

environmentally persistent free radicals at a superfund wood treating site. Environ. Sci. Technol. 2011, 45 (15), 6356-6365. (6) Govindaraj, N.; Mortland, M. M.; Boyd, S. A. Single electron-transfer mechanism of oxidative dechlorination of 4-chloroanisole on copper(II)-smectite. Environ. Sci. Technol. 1987, 21 (11), 1119-1123. (7) Boyd, S. A.; Mortland, M. M. Radical formation and polymerization of chlorophenols and chloroanisole on copper(II)-smectite. Environ. Sci. Technol. 1986, 20 (10), 1056-1058. (8) Pandey, A. K.; Pandey, S. D.; Misra, V.; Viswanathan, P. N. Role of free radicals in the binding of organochlorine pesticides and heavy metals with humic acid. Sci. Total Environ. 1999, 231 (2-3), 125-133. (9) Fingler, S.; Drevenkar, V.; Frobe, Z. Sorption of chlorophenolates in soils and aquifer and marine sediments. Arch. Environ. Contam. Toxicol. 2005, 48 (1), 32-39. (10) Sawhney, B. L.; Reynolds, R. C. Interstratified clays as fundamental particles - a discussion. Clay Clay Miner. 1985, 33 (6), 559-559. (11) Jia, H. Z.; Li, L.; Chen, H. X.; Zhao, Y.; Li, X. Y.; Wang, C. Y. Exchangeable cations-mediated photodegradation of polycyclic aromatic hydrocarbons (PAHs) on smectite surface under visible light. J. Hazard. Mater. 2015, 287, 16-23. (12) Zhang, W. H.; Zhuang, L. W.; Yuan, Y. A.; Tong, L. Z.; Tsang, D. C. W. Enhancement of phenanthrene adsorption on a clayey soil and clay minerals by coexisting lead or cadmium. Chemosphere 2011, 83 (3), 302-310. (13) Xiao, L.; Qu, X. L.; Zhu, D. Q. Biosorption of nonpolar hydrophobic organic compounds to Escherichia coli facilitated by metal and proton surface binding. Environ. Sci. Technol. 2007, 41 (8), 2750-2755. (14) Tamamura, S.; Sato, T.; Ota, Y.; Tang, N.; Hayakawa, K. Decomposition of polycyclic aromatic hydrocarbon (PAHs) on mineral surface under controlled relative humidity. Acta Geol. Sin-Engl. 2006, 80 (2), 185-191. (15) Zhang, W.; Zhuang, L.; Yuan, Y.; Tong, L.; Tsang, D. C. Enhancement of phenanthrene adsorption on a clayey soil and clay minerals by coexisting lead or cadmium. Chemosphere 2011, 83 (3), 302-10. (16) Boyd, S. A.; Mortland, M. M. Dioxin radical formation and polymerization on Cu(II)-smectite. Nature 1985, 316 (6028), 532-535. (17) Soma, Y.; Soma, M.; Harada, I. Reactions of aromatic-molecules in the interlayer of transition-metal ion-exchanged montmorillonite studied by resonance raman-spectroscopy .2. monosubstituted benzenes and 4,4'-disubstituted biphenyls. J. Phys. Chem-Us. 1985, 89 (5), 738-742. (18) Jia, H. Z.; Zhao, J. C.; Li, L.; Li, X. Y.; Wang, C. Y. Transformation of polycyclic aromatic hydrocarbons (PAHs) on Fe(III)-modified clay minerals: Role of molecular chemistry and clay surface properties. Appl. Catal. B-Environ. 2014, 154, 238-245. (19) Qu, X.; Wang, X.; Zhu, D. The partitioning of PAHs to egg phospholipids facilitated by copper and proton binding via cation-pi interactions. Environ. Sci. Technol. 2007, 41 (24), 8321-7. (20) Suzuki, S.; Green, P. G.; Bumgarner, R. E.; Dasgupta, S.; Goddard, W. A.; Blake, G. A. Benzene forms hydrogen-bonds with water. Science 1992, 257 (5072), 942-944. (21) Kumpf, R. A.; Dougherty, D. A. A mechanism for ion selectivity in potassium channels: computational studies of cation-pi interactions. Science 1993, 261 (5129), 1708-1710. (22) Zhu, D.; Herbert, B. E.; Schlautman, M. A.; Carraway, E. R.; Hur, J. Cation-pi bonding: a new 24


Page 25 of 32


561 562 563 564 565 566 567 568 569 570 571 572 573 574 575 576 577 578 579 580 581 582 583 584 585 586 587 588 589 590 591 592 593 594 595 596 597 598 599 600 601 602 603 604

perspective on the sorption of polycyclic aromatic hydrocarbons to mineral surfaces. J.Environ. Qual. 2004, 33 (4), 1322-30. (23) Goss, K. U.; Schwarzenbach, R. P. Linear free energy relationships used to evaluate equilibrium partitioning of organic compounds. Environ. Sci. Technol. 2001, 35 (1), 1-9. (24) Gokel, G. W.; Barbour, L. J.; De Wall, S. L.; Meadows, E. S. Macrocyclic polyethers as probes to assess and understand alkali metal cation-pi interactions. Coordin. Chem. Rev. 2001, 222, 127-154. (25) Qu, X. L.; Liu, P.; Zhu, D. Q. Enhanced sorption of polycyclic aromatic hydrocarbons to tetra-alkyl ammonium modified smectites via cation-pi interactions. Environ. Sci. Technol. 2008, 42 (4), 1109-1116. (26) Dellinger, B.; Loninicki, S.; Khachatryan, L.; Maskos, Z.; Hall, R. W.; Adounkpe, J.; McFerrin, C.; Truong, H. Formation and stabilization of persistent free radicals. P. Combust. Inst. 2007, 31, 521-528. (27) Valavanidis, A.; Iopoulos, N.; Gotsis, G.; Fiotakis, K. Persistent free radicals, heavy metals and PAHs generated in particulate soot emissions and residue ash from controlled combustion of common types of plastic. J. Hazard. Mater. 2008, 156 (1-3), 277-284. (28) Jia, H. Z.; Zhao, J. C.; Fan, X. Y.; Dilimulati, K.; Wang, C. Y. Photodegradation of phenanthrene on cation-modified clays under visible light. Appl. Catal. B-Environ. 2012, 123, 43-51. (29) Delley, B. An all-electron numerical-method for solving the local density functional for polyatomic-molecules. J. Chem. Phys. 1990, 92 (1), 508-517. (30) Delley, B. From molecules to solids with the DMol(3) approach. J. Chem. Phys. 2000, 113 (18), 7756-7764. (31) Becke, A. D. A multicenter numerical-integration scheme for polyatomic-molecules. J. Chem. Phys. 1988, 88 (4), 2547-2553. (32) Perdew, J. P.; Wang, Y. Accurate and simple analytic representation of the electron-gas correlation-energy. Phys. Rev. B 1992, 45 (23), 13244-13249. (33) Lehner, A. F.; Horn, J.; Flesher, J. W. Formation of radical cations in a model for the metabolism of aromatic hydrocarbons. Biochem. Bioph. Res. Co. 2004, 322 (3), 1018-1023. (34) Di Valentin, C.; Neyman, K. M.; Risse, T.; Sterrer, M.; Fischbach, E.; Freund, H. J.; Nasluzov, V. A.; Pacchioni, G.; Rosch, N. Density-functional model cluster studies of EPR g tensors of F-s(+) centers on the surface of MgO. J. Chem. Phys. 2006, 124 (4), 044708. (35) Hales, B. J.; Case, E. E. Immobilized radicals .4. Biological semi-quinone anions and neutral semiquinones. Biochim. Bioph. Acta 1981, 637 (2), 291-302. (36) Qin, C.; Troya, D.; Shang, C.; Hildreth, S.; Helm, R.; Xia, K. Surface catalyzed oxidative oligomerization of 17 beta-estradiol by Fe(III)-saturated montmorillonite. Environ. Sci. Technol. 2015, 49 (2), 956-964. (37) Eastman, M. P.; Patterson, D. E.; Pannell, K. H. Reaction of benzene with Cu(II)-exchanged and Fe(III)-exchanged hectorites. Clay Clay Miner. 1984, 32 (4), 327-333. (38) Jezierski, A.; Skrzypek, G.; Jezierski, P.; Paul, D.; Jedrysek, M. O. Electron paramagnetic resonance (EPR) and stable isotope records of paleoenvironmental conditions during peat formation. Spectrochim. Acta A 2008, 69 (5), 1311-1316. (39) Christoforidis, K. C.; Un, S.; Deligiannakis, Y. High-field 285 GHz electron paramagnetic resonance study of indigenous radicals of humic acids. J. Phys. Chem. A 2007, 111 (46), 11860-11866. (40) Khachatryan, L.; Adounkpe, J.; Maskos, Z.; Dellinger, B. Formation of cyclopentadienyl radical from the gas-phase pyrolysis of hydroquinone, catechol, and phenol. Environ. Sci. Technol. 2006, 40 25



605 606 607 608 609 610 611 612 613 614 615 616 617 618 619 620 621 622 623 624 625 626 627 628 629 630 631 632 633 634 635 636 637 638 639 640 641 642 643 644 645 646 647 648

Page 26 of 32

(16), 5071-5076. (41) Cheng, F. C.; Jen, J. F.; Tsai, T. H. Hydroxyl radical in living systems and its separation methods. J. Chromatogr. B 2002, 781 (1-2), 481-496. (42) Huang, X.; Zalma, R.; Pezerat, H. Chemical reactivity of the carbon-centered free radicals and ferrous iron in coals: role of bioavailable Fe(II) in coal workers pneumoconiosis. Free Radic. Res. 1999, 30 (6), 439-51. (43) Si, Y. B.; Wang, S. Q.; Zhou, D. M.; Chen, H. M. Adsorption and photo-reactivity of bensulfuron-methyl on homoionic clays. Clay Clay Miner. 2004, 52 (6), 742-748. (44) Gu, C.; Liu, C.; Ding, Y. J.; Li, H.; Teppen, B. J.; Johnston, C. T.; Boyd, S. A. Clay Mediated Route to Natural Formation of Polychlorodibenzo-p-dioxins. Environ. Sci. Technol. 2011, 45 (8), 3445-3451. (45) Gu, C.; Li, H.; Teppen, B. J.; Boyd, S. A. Octachlorodibenzodioxin formation on Fe(III)-montmorillonite clay. Environ. Sci. Technol. 2008, 42 (13), 4758-4763. (46) Vejerano, E.; Lomnicki, S. M.; Dellinger, B. Formation and stabilization of combustion-generated, environmentally persistent radicals on Ni(II)O supported on a silica surface. Environ. Sci. Technol. 2012, 46 (17), 9406-9411. (47) Li, H.; Pan, B.; Liao, S. H.; Zhang, D.; Xing, B. S. Formation of environmentally persistent free radicals as the mechanism for reduced catechol degradation on hematite-silica surface under UV irradiation. Environ. Pollut. 2014, 188, 153-158. (48) Barckholtz, C.; Fadden, M. J.; Hadad, C. M. Computational study of the mechanisms for the reaction of O-2((3)Sigma(g)) with aromatic radicals. J. Phys. Chem. A 1999, 103 (40), 8108-8117. (49) dela Cruz, A. L. N.; Cook, R. L.; Lomnicki, S. M.; Dellinger, B. Effect of low temperature thermal treatment on soils contaminated with pentachlorophenol and environmentally persistent free radicals. Environ. Sci. Technol. 2012, 46 (11), 5971-5978. (50) Rupert, J. P. Electron-spin resonance spectra of interlamellar Cu(II)-arene complexes on montmorillonite. J. Phys. Chem. 1973, 77 (6), 784-790. (51) Johnston, C. T.; Tipton, T.; Trabue, S. L.; Erickson, C.; Stone, D. A. Vapor-phase sorption of p-xylene on Co-exchanged and Cu-exchanged Saz-1 montmorillonite. Environ. Sci. Technol. 1992, 26 (2), 382-390. (52) Polubesova, T.; Eldad, S.; Chefetz, B. Adsorption and oxidative transformation of phenolic acids by Fe(III)-montmorillonite. Environ. Sci. Technol. 2010, 44 (11), 4203-4209. (53) Liyanapatirana, C.; Gwaltney, S. R.; Xia, K. Transformation of triclosan by Fe(III)-saturated montmorillonite. Environ. Sci. Technol. 2010, 44 (2), 668-674. (54) Sorokin, A.; Meunier, B. Oxidation of polycyclic aromatic hydrocarbons catalyzed by iron tetrasulfophthalocyanine FePcS: Inverse isotope effects and oxygen labeling studies. Eur. J. Inorg. Chem. 1998, 9, 1269-1281. (55) Wang, D. G.; Li, Y. F.; Yang, M.; Han, M. Decomposition of polycyclic aromatic hydrocarbons in atmospheric aqueous droplets through sulfate anion radicals: An experimental and theoretical study. Sci. Total Environ. 2008, 393 (1), 64-71. (56) Samanta, S. K.; Singh, O. V.; Jain, R. K. Polycyclic aromatic hydrocarbons: environmental pollution and bioremediation. Trend. Biotechnol. 2002, 20 (6), 243-248. (57) dela Cruz, A. L.; Cook, R. L.; Dellinger, B.; Lomnicki, S. M.; Donnelly, K. C.; Kelley, M. A.; Cosgriff, D. Assessment of environmentally persistent free radicals in soils and sediments from three Superfund sites. Environ. Sci. Proc. Impacts 2014, 16 (1), 44-52. 26


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(58) Vejerano, E.; Lomnicki, S.; Dellinger, B. Formation and stabilization of combustion-generated environmentally persistent free radicals on an Fe2O3/silica surface. Environ. Sci. Technol. 2011, 45 (2), 589-594. (59) Boyd, S. A. M., M. M. Radical formation and polymerization of chlorophenols and chloroanisole on copper(II)-smectite. Environ. Sci. Technol. 1986, 20, 1056-1058. (60) Balakrishna, S.; Lomnicki, S.; McAvey, K. M.; Cole, R. B.; Dellinger, B.; Cormier, S. A. Environmentally persistent free radicals amplify ultrafine particle mediated cellular oxidative stress and cytotoxicity. Part. Fibre. Toxicol. 2009, 6 (2), 139-145.

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Figure Captions 1.5

80

2.0038

g-Factor

0.5 0.0

8d 10d 12d 17d 25d

-0.5 -1.0

100

b

1d 2d 3d 4d 7d

1.0

Intensity (a.u.)

2.0040

a

g-factor Peak area

40

2.0034

-1.5

20

2.0032 1.99

2.00

2.01

0

0

2.02

5

10

g-Factor

g1 (2.0028-2.0030) g2 (2.0036-2.0038) g3 (2.0031-2.0032)

25

30

35

d k3 = 0.026 t1/e= 38.46 d

-0.5 ln(C/C0)

80

20

0.0

c 100

15

Time (d)

120

Peak area

60

2.0036

Peak area

675

Page 28 of 32

60 40

-1.0 g1 g2 g3

-1.5

20 -2.0 0 -2.5 0

5

10

15

20

25

30

35

k2 = 0.29 t1/e= 3.45 d

k1 = 0.71 t1/e= 1.41 d

0

2

4

6

676

8

10

12

14

16

Time (d)

Time(d)

677

Fig. 1. The evolution of (a) EPR spectra and (b) their g-factor and peak area as a

678

function of reaction time in the reaction system of anthracene-contaminated

679

Fe(III)-montmorillonite at relative humidity (RH) of 8% and room temperature (~ 25

680

o

681

time. (d) Normalized pseudo-first-order decay kinetics of EPFRs derived from the

682

reaction time at their highest yield for various radicals, i.e., 3d for g1, 10d for g2, and

683

15d for g3.

C). (c) The evolution of peak area of g1, g2, and g3 signals as function of reaction

684 685 686 687 688 689 690 691 692

28


Page 29 of 32


2.0044

70 In glovebox

2.0042

In air 60

2.0040 50

2.0036 40 2.0034 2.0032

Peak area

g-Factor

2.0038

30

2.0030 20

g-factor Peak area

2.0028 2.0026

10 0

10

20

30

40

50

Time (d) 693 694

Fig. 2. The evolution of g-factor and peak area of EPR spectra as a function of

695

reaction time in the reaction system of anthracene-contaminated Fe(III)-

696

montmorillonite under anoxic and oxic conditions at room temperature (~ 25 oC).

697 698 699 700 701 702 703 704 705 706 707 708 709 710 711 712 713 714 715 716 717 718

29



Intensity (a.u.)

0.6 0.4 0.2 0.0

60 2.00360

a

g-factor Peak area

2.00355

-0.2

b

50 40

2.00350 2.00345

30

2.00340

20

2.00335

10

Peak area

25 oC o 40 C 50 oC 60 oC 70 oC 80 oC 90 oC

g-Factor

0.8

Page 30 of 32

-0.4 -0.6 -0.8

2.00330

0 20

1.990 1.995 2.000 2.005 2.010 2.015

1.0

25 oC 40 oC

80

100

0.8

50 oC 60 oC

0.6

70 oC 90 oC

1.0

c

d 0.8

C/C0

C/C0

60

Temprature (oC)

g-Factor

0.4

0.6 RH = 8% RH = 11% RH = 30% RH = 60% RH = 90%

0.4

0.2

0.2

0.0

0.0 0

719

40

20

40

60

80

0

20

Time (h)

40

60

80

Time (h)

720

Fig. 3. (a) EPR spectra under different reaction temperatures, and (b) Temperature

721

dependence of g-factors and peak area in the 15 d reaction system of

722

anthracene-contaminated Fe(III)-montmorillonite. The transformation of anthracene

723

as a function of reaction time at Fe(III)-montmorillonite surface under various

724

reaction temperatures (c) and relative humidity (d).

725 726 727 728 729 730 731 732 733 734 735 736 737 738 739 30


Page 31 of 32


Relative energy (kcal/mol)

1.945 TS + O2 0.355 IM1

IM2

0.142

+ H2O 0.260

TS

.

+

TS

Radical A

-13.673

-22.591 Radical B -30.517

740 741

Fig. 4. Profile of the reaction of anthracene with Fe(III). The energies of anthracene

742

complexes with Fe(III) were set to zero.

743 744 745 746 747 748 749 750 751 752 753 754 755 756 757 758 759 760 761 762 763

31



H Fe(III)O+

-H + H2 O

Radical A

H

OH

OH Fe(II)OH+ Fe(III)(H2O)

Fe(II)O Fe(II)OH+

Fe(II)O

Page 32 of 32

-H

Radical B

OH

OH

Fast

Intermediate B

Intermediate A

O

O

O

O Fe(II)O Fe(III)O+

Fe(III)(H2O) Fe(II)O

-H + H2 O

764

O

H

OH

Radical C

765

Scheme 1. Proposed mechanism for the transformation of anthracene and formation of

766

EPFRs on Fe(III)-modified montmorillonite.

767 768

32


Formation and Stabilization of Environmentally Persistent Free

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