Formation and Stabilization of Environmentally Persistent Free

May 25, 2016 - Environmentally persistent free radicals (EPFRs) are occasionally detected in Superfund sites but the formation of EPFRs induced by pol...
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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

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Formation and stabilization of environmentally persistent free

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radicals induced by the interaction of anthracene with

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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

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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)

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ABSTRACT:

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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

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formation of EPFRs on anthracene-contaminated clay minerals was quantitatively

31

monitored

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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

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

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Transformation pathway and possible mechanism are proposed on the basis of

41

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

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

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

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

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toxicity than PAHs and exerts significant impacts on human health.

46

KEYWORD:

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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

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TOC Art

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INTRODUCTION

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

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

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

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

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previously observed in combustion-generated particles and airborne particulate matter

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

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

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

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

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persist in ambient air for days.3,

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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

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found in Superfund sites contaminated with polycyclic aromatic hydrocarbons

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

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

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previously suspected or envisioned, especially at sites contaminated with organic

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pollutants.

4

Recently, EPFRs were detected in

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

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

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

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

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

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

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

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

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

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

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

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

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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

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

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electron transfer, often followed by further reactions including dechlorination and

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polymerization.6,

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through the formation of organic radicals as intermediates with simultaneous

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

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

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

7, 16, 17

Transformation of these pollutants is possibly achieved

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environment, i.e., EPFRs.4

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

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

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

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

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interaction has been demonstrated as an important factor regulating PAHs availability

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

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

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

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

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

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

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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

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

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

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

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soil in association with EPFRs.

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

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

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environmentally relevant conditions. Conversion of molecular PAHs to EPFRs as

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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

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interactions between PAHs and clay minerals in PAHs transformation and EPFRs

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

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

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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

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PAHs-contaminated clay minerals in association with free radicals.

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EXPERIMENTAL SECTION

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

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

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

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

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interlayer swelling properties (Table S1).

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

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

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

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

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

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

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

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

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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

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

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ambient conditions.

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

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

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

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

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

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

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

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

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

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

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and 5 mL dichloromethane) and analyzed immediately.

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EPR Characterization. All EPR measurements were performed using a Bruker

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E500 EPR spectrometer at room temperature. Instrument and operating parameters

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

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

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

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

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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

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

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

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

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

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

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

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

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

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o

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

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

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

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

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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)

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

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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:

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Ln(R / R0 ) = −kt , and t1/e = 1/ k

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

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

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

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

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

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

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and transition state structures were performed using the Becke exchange31 plus

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Perdew-Wang approximation functional32 within Generalized gradient approximation

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

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located by performing relaxed potential energy surface scans followed by

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

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synchronous transit (QST) method.

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RESULTS AND DISCUSSION

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

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

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

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

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

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systems, significant EPR signals were observed in anthracene-contaminated

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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).

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

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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

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

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

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

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during the experimental period. For control experiments performed with

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

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

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

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

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

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

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the

217

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

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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

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

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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

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

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

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

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

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

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

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

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

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

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

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

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37

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

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

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

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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

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

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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

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

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

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

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

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

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

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

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

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

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

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

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

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transformation products of anthracene could be identified as anthraquinone formed by

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

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considered as a major detectable intermediate for the transformation from anthracene

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

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

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

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

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

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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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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thereby making them stable and persistent.41,

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

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Fe(III)-kaolinite,

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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

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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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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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,

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

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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

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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

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

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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

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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,

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

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

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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

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

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environmentally persistent.

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Effect of environmental condition

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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

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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

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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

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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

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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

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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

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421

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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

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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

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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

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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

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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.

657 658 659 660 661 662 663 664 665 666 667 668 669 670 671 672 673 674 27

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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

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Page 29 of 32

Environmental Science & Technology

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

ACS Paragon Plus Environment

Environmental Science & Technology

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

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Page 31 of 32

Environmental Science & Technology

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

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ACS Paragon Plus Environment

Environmental Science & Technology

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

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