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Rapid Selective Circumneutral Degradation of Phenolic Pollutants Using Peroxymonosulfate-Iodide Metal-Free Oxidation: Role of Iodine Atoms Yong Feng, Po Heng Lee, Deli Wu, and Kaimin Shih Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b04528 • Publication Date (Web): 27 Jan 2017 Downloaded from http://pubs.acs.org on January 28, 2017

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Environmental Science & Technology

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Rapid Selective Circumneutral Degradation of Phenolic Pollutants Using

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Peroxymonosulfate-Iodide Metal-Free Oxidation: Role of Iodine Atoms

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Yong Fenga, Po-Heng Leeb, Deli Wuc, Kaimin Shiha,*

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a

Department of Civil Engineering, The University of Hong Kong, Pokfulam, Hong Kong Department of Civil and Environmental Engineering, The Hong Kong Polytechnic University, Hung Hom, Hong Kong c State Key Laboratory of Pollution Control and Resources Reuse, School of Environmental Science & Engineering, Tongji University, Shanghai 200092, People’s Republic of China b

Contact information:

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Yong Feng ([email protected])

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Po-Heng Lee ([email protected])

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Deli Wu ([email protected])

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Kaimin Shih ([email protected])

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Manuscript to be submitted to Environmental Science & Technology

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* Corresponding author:

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Dr. Kaimin Shih

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Phone: +852-2859-1973

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Fax: +852-2559-5337

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E-mail: [email protected]

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WORD COUNTS

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Text about 5620 words (Abstract, Manuscript Body, Acknowledgements, and Description of

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Supporting Information) + 1500 (5 Figures) = 7120 words

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ABSTRACT

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The development of environmentally friendly, oxidation-selective advanced oxidation processes

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(AOPs) for water decontamination is important for resource recovery, carbon dioxide abatement,

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and cost savings. In this study, we developed an innovative AOP using a combination of

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peroxymonosulfate (PMS) and iodide ions (I–) for the selective removal of phenolic pollutants

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from aqueous solutions. The results showed that nearly 100% degradation of phenol, bisphenol

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A, and hydroquinone was achieved after reacting for 4 min in the presence of 65 µM PMS and

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50 µM I–. PMS-I– oxidation had a wide effective pH range, with the best performance achieved

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under circumneutral conditions. The ratio between [PMS] and [I–] influenced the degradation,

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and the optimal ratio was approximately 1.00 for the degradation of the phenols. Neither sulfate

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nor hydroxyl radicals were found to be the active species in PMS-I– oxidation. Instead, we found

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evidence that iodide atoms were the dominant oxidants. In addition, both Cl– and Br– also

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promoted the degradation of phenol in PMS solution. The results of this work may promote the

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application of reactive halogen species in water treatment.

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INTRODUCTION

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Hydroxyl radical (·OH)–based advanced oxidation processes (HR-AOPs) have been extensively

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studied for the pollution abatement of wastewater that contains refractory contaminants.1,

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However, one major disadvantage associated with HR-AOPs is that ·OH is a nonselective

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oxidant.3,

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controlled rates5 and can be easily scavenged by various aqueous components, such as

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bicarbonate,6 chloride,7 and natural organic matter.8 In addition, some compounds that are

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readily biodegradable are also mineralized in HR-AOPs when used as pretreatments,4 which

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inevitably consume a large fraction of the radicals, together with peroxides or the energy

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required in these processes. Therefore, the development of AOPs with high selectivity toward

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specific pollutants is a promising approach to reduce costs and expand the effective application

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of AOPs in the wastewater treatment industry.

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·OH reacts with common organic and inorganic compounds at nearly diffusion-

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Persulfates, including peroxymonosulfate (PMS, HSO ) and peroxydisulfate (PDS, S O ),

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have attracted considerable attention for the production of sulfate radicals (SO∙ ) in the past

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decade.9-11 Importantly, some recent investigations have shown that persulfate can be activated

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by a non-radical approach; pollutants were degraded by donating electrons to PMS or PDS at

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active sites, in which the catalysts functioned as electron shuttles. Zhang et al.12 investigated the

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combination of PDS/CuO for 2,4-dichlorophenol degradation and found that no SO∙ was

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produced by this combination. Instead, indirect evidence suggested that a non-radical mechanism

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via outer-sphere interactions may dominate the degradation. Duan et al.13 studied phenol

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degradation by PMS under the catalysis of N-doped carbon nanotubes. Ethanol, a highly efficient

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scavenger of SO∙ , had little influence on the degradation of phenol. Instead, the degradation



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performance was positively related to the content of the N dopant, which served to promote

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electron transfer in the carbon nanotubes.

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Recently, reactive halogen species (RHS), including halogen atoms (X·) and halogen radical

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14-16 anions ( X ∙ ), usually produced during the scavenging of ·OH or SO∙ have by halides,

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received increasing attention in the development of water purification technologies due to their

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relatively strong oxidation capability,15 their greater selectivity than ·OH,17 and the ubiquitous

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presence of halide ions in water bodies (Table S1). In contrast to ·OH, which almost exclusively

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oxidizes compounds via H-abstraction and addition, RHS mainly reacts with pollutants via one-

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electron oxidation,15,

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addition, the formation of halogenated byproducts has been found to be minimal when phenol

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was degraded by HR-AOPs in the presence of both Cl– and Br– ions.17 Furthermore, RHS are

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only weakly reactive with carboxylic acids and alcohols, which are the most common

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degradation products of organic contaminants in AOPs.19 For example, Cl∙ reacts with methanol

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and acetic acid at rate constants of only 3.5 × 103 M–1 s–1 and < 104 M–1 s–1,14 respectively,

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suggesting its potential utility for the development of product-oriented AOPs for wastewater

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treatment.

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although H-abstraction and addition have also been reported.14,

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In

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As one of the common halides, iodide (I–) is essential to human bodily health for the

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synthesis of thyroid hormones and is ubiquitous in aquatic environments (Table S1). Unlike Cl–

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and Br–, for which the generation of hazardous chlorate and bromate is of concern, the

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production of iodate by the oxidation of I– has been considered as a desirable route because

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iodate can be endogenically transformed back to I–.20 In this work, we proposed an innovative

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metal-free AOP for the rapid selective degradation of phenols using a novel combination of PMS

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and I–. Structurally diverse sets of representative phenolic compounds (phenol, bisphenol A, 4 ACS Paragon Plus Environment

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hydroquinone, and 2,4-dichlorophenol) and non-phenolic compounds (nitrobenzene, atrazine,

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chloramphenicol, and methylene blue; Table S2) were used to evaluate the capability and

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selectivity of the combination of PMS and I–. The degradation of the phenolic pollutants by

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PMS-I– oxidation was systematically studied, and a mechanism involving iodide atoms (I·) as the

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dominant oxidants was proposed. Finally, the potential of the PMS-Cl– and PMS-Br– systems to

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produce RHS was also explored.

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EXPERIMENTAL

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Chemicals and Materials. Bisphenol A (≥99%), nitrobenzene (≥99%), 2,4-dichlorophenol

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(99%), Oxone (KHSO5·½KHSO4·½K2SO4), sodium chloride (≥99.5%), potassium bromide

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(99.95%), resorcinol (≥99%), sodium bicarbonate (99.5% to 100.5%), and dichloromethane

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(99.9%), and ammonium acetate (liquid chromatography-mass spectrometry (LC-MS) ultra)

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were purchased from Sigma-Aldrich (St. Louis, MO). Hydroquinone (≥99%), phenol (≥99%),

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potassium iodide (99.8%), and methylene blue were obtained from BDH Chemicals (Poole, UK).

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Chloramphenicol (98%) and atrazine (97%) were supplied by J&K Chemicals (Hong Kong) and

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TCI (Shanghai, China), respectively. 2-iodophenol (98%), 3-iodophenol (98%), and 2,4,6-

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triiodophenol (98%) were purchased from Alfa Aesar (Heysham, UK). Sodium sulfate (99.99%)

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was purchased from Merck (Darmstadt, Germany). Methanol (Optima LC-MS grade) and

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sodium sulfite (≥98%) were obtained from Fisher Scientific (Pittsburgh, PA). Other chemicals

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were of ACS reagent grade or higher and were used as received without further purification. The

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sodium chloride obtained from Sigma-Aldrich was reported to have < 0.001% I– and < 0.01%

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Br– by weight. All solutions were prepared with ultrapure water (18.0 MΩ · cm) from a

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Barnstead EASYpure UV/UF purification system (Dubuque, IA). 5 ACS Paragon Plus Environment


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Experimental Procedures. All experiments were conducted in 50-mL polypropylene centrifuge

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tubes (Bio-Rad Laboratories, Hercules, CA) at room temperature (23 ± 2 °C). First, 40-mL

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solutions with a specified pollutant concentration and an ionic strength provided by 10 mM

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Na2SO4 (further specified in Text S1) were transferred to the tubes. Their pH values were

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adjusted when necessary using 0.05 M NaOH or H2SO4. Due to only a slight variation in the

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solution pH during the reaction and the insignificant effect of phosphate buffer (Figure S1),

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buffer solutions were not used when evaluating the capability of PMS-I– oxidation. A specified

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amount of PMS stock solution (66 mM) was then spiked into the tube reactors, followed by the

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addition of KI stock solution (100 mM) to initiate the reaction. The resulting solutions were

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immediately mixed by agitation on a VX-100 vortex mixer (Labnet, Edison, NJ) for

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approximately 2 s. Samples (1 mL) were withdrawn with a 1-mL pipette at specified time

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intervals and transferred to 2-mL LC vials (MACHEREY-NAGEL, Düren, Germany) for

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analysis. 50 µL sodium sulfite (2 M), reacting rapidly with I· ([1.0 ± 0.3] × 109 M–1 s–1)21 and

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other oxidants (Table S3), was immediately added to the vials. To mitigate any potential

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degradation, we analyzed each sample within 20 min after withdrawal. For total organic carbon

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(TOC) analysis, the reactions were quenched after 10 min with excess sodium sulfite powder. All

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degradation experiments were performed in duplicate or triplicate. The data obtained were

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averaged, and the corresponding standard deviation was presented.

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Chemical Analysis. Bisphenol A, nitrobenzene, atrazine, chloramphenicol, phenol, 2,4-

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dichlorophenol, and hydroquinone were analyzed with a Waters AQUITY Ultra Performance LC

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(UPLC) system equipped with a photodiode array (PDA) detector. A Waters BEH C18 column 6 ACS Paragon Plus Environment

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(50 mm × 2.1 mm, 1.7 µm) with a VanGuard pre-column (5 mm × 2.1 mm, 1.7 µm) was used for

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the separation. Mixtures containing varied ratios of ultrapure water and Optima LC-MS grade

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methanol were used as the mobile phase for these compounds. Detailed analytical parameters

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and the calibration ranges are listed in Text S2 and Table S4. Methylene blue was analyzed

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spectrometrically at 670 nm with a Biochrom Libra S12 UV-visible spectrometer (Wolf

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Laboratories, Pocklington, UK). TOC was determined with a Shimadzu TOC-V CPH analyzer

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(Kyoto, Japan) using the 680 °C combustion method. The degradation products of phenol were

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first identified with an Agilent gas chromatography (GC)-MS (6890N-5973) and then quantified

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with the UPLC system connected to a Waters electrospray ionization-triple quadrupole mass

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spectrometer (MS/MS) in multiple-reaction-monitoring modes. Details are listed in Text S3. The

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pH values were analyzed with an Orion 2-Star benchtop pH meter.

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Identification of Active Species. Due to the asymmetrical structure of PMS, both SO∙ and ·OH

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can be formed by PMS-I– oxidation. To evaluate the contribution of these radicals to the

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degradation, the effect of methanol, a well-known, efficient scavenger for both SO∙ and ·OH, on

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the degradation of the phenols was explored. In addition, formaldehyde, the product of methanol

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oxidation by these radicals, was quantified after derivatization with 2,4-dinitrophenylhydrazine

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(DNPH)22 to further examine the production of radicals. Details are listed in Text S4. To

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evaluate the role of I· in the degradation of the phenols, we spectrometrically monitored the

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production of I , the reaction product of I· and I–, at 352 nm.

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RESULTS AND DISCUSSION

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Degradation of Selected Pollutants by PMS-I– Oxidation. Although PMS is a strong oxidant,

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when used in isolation it showed a negligible influence on the concentration of the selected

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pollutants, except for hydroquinone (Figure S2), in the tested time course. The degradation of

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hydroquinone by PMS alone was probably due to the co-presence of 1,4-benzoquinone, which

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was produced by the oxidation of hydroquinone in air. 1,4-benzoquinone can efficiently activate

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PMS to produce singlet oxygen (1O2),23 an oxidant that may have contributed to the degradation

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of the pollutants. The degradation of the various pollutants, including bisphenol A, nitrobenzene,

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methylene blue, atrazine, chloramphenicol, phenol, 2,4-dichlorophenol, and hydroquinone, by

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PMS-I– oxidation is shown in Figure 1a. In the presence of 65 µM PMS and 50 µM I–, the

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concentrations of nitrobenzene, atrazine, chloramphenicol, and methylene blue showed no

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obvious decline, suggesting that PMS-I– oxidation was inefficient for these pollutants. However,

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under identical conditions, the degradation of bisphenol A, hydroquinone, and phenol occurred

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rapidly; nearly 100% degradation of these pollutants was realized within 4 min. 2,4-

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dichlorophenol was relatively resistant to degradation, with only ~61.5% degradation within the

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same time course, but this increased to ~89% after 10 min. PMS alone showed no ability to

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degrade phenols except for hydroquinone (Figure S2). Even in the case of hydroquinone, its

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degradation rate by PMS-I– oxidation was nearly 14 times that with PMS alone after 2 min.

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Therefore, the rapid degradation of these phenols was evidently related to the interactions

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between PMS and I–. Because bisphenol A, phenol, hydroquinone, and 2,4-dichlorophenol are

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phenolic compounds and atrazine, nitrobenzene, chloramphenicol, and methylene blue are non-

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phenolic compounds, the difference in their degradation rates suggests that PMS-I– oxidation has

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a selective capability toward phenolic pollutants. This capability was further demonstrated by the


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selective degradation of phenol when phenol co-existed with atrazine in PMS-I– oxidation

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(Figure S3). Kinetic investigation showed that the degradation of the phenols followed pseudo-

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first-order kinetics, with constant kapp values of 1.01, 0.22, and 1.30 min–1 for phenol, 2,4-

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dichlorophenol, and bisphenol A, respectively (Figure S4). Due to the speed of the reaction, the

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kinetic constant of hydroquinone degradation could not be calculated, but is expected to be

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significantly greater than that of bisphenol A (Figure 1a). To further explore the degradation

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capability of PMS-I– oxidation, the TOC removal of phenol and 2,4-dichlorophenol was

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examined. To facilitate the analysis, the initial concentrations of these two pollutants were

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increased to 100 and 62 µM, respectively. As shown in Figure 1b, removal of approximately 61%

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and 6% of TOC from the phenol and 2,4-dichlorophenol solutions, respectively, were achieved

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after 10 min in the presence of 325 µM PMS and 325 µM I–. The successful removal of TOC

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suggests that PMS-I– oxidation had the capability to mineralize the phenolic pollutants, although

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we did not optimize the experimental parameters to maximize the TOC removal.

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Figure 1. Degradation of selected pollutants by PMS-I– oxidation (a) and TOC removal of phenol

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and 2,4-dichlorophenol by PMS-I– oxidation with different doses of I–. Conditions:

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[hydroquinone] = [bisphenol A] = [nitrobenzene] = [chloramphenicol] = [methylene blue] = 10



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µM (a), [phenol] = 10 µM (a) or 100 µM (b), [2,4-diclorophenol] = 10 µM (a) or 62 µM (b),

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[PMS] = 65 µM (a) or 325 µM (b), [I–] = 50 µM (a), reaction time = 10 min (b), [Na2SO4] = 10

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mM, and initial solution pH = 6.0.

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Dependence of Oxidation on PMS/I– Ratio and Initial Solution pH Values. The effect of the

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ratio between [PMS] and [I–] on phenol and 2,4-diclorophenol degradation was investigated by

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examining various doses of I– (Figures 2a to 2d). When [PMS]/[I–] was decreased from 6.50 to

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1.30, an increase in the degradation of phenol was observed (Figure 2a). However, a further

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decrease in the ratio from 1.00 to 0.13 slowed down the degradation; approximately 90% of the

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phenol was degraded with the [PMS]/[I–] ratio of 1.30, whereas only about 60% of the phenol

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was degraded when the ratio was reduced to 0.13. These results suggest the existence of an

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optimal ratio between [PMS] and [I–]. The pseudo-first-order constants calculated with different

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[PMS]/[I–] ratios (Figure 2b) were consistent with the degradation of phenol; a plateau was

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reached when [PMS]/[I–] was fixed at 1.00. Similar effects of the [PMS]/[I–] ratio on the

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degradation of 2,4-diclorophenol were also observed (Figures 2c and 2d), and the optimal ratio

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was consistently found to be approximately 1.00.

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The initial pH value of the solution also influenced the degradation of the phenols by PMS-

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I– oxidation (Figure 2e). Strongly basic conditions were not beneficial for degradation; less than

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5% of the phenol was degraded at pH 11.5 after 10 min with 65 µM PMS and 20 µM I– ions,

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whereas overall degradations of ~59%, 66%, 62%, and 56% were achieved with initial solution

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pH values of 4.0, 6.0, 8.0, and 10.0, respectively. Due to the dissolution of protons released from

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Oxone (KHSO₅·½KHSO₄·½K₂SO₄), the actual pH values after adding PMS changed from 4.0,

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6.0, 8.0, and 10.0 to around 3.8, 4.5, 4.9, and 9.5, respectively. On the basis of these actual pH

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values and the corresponding degradation performances, it can be concluded that strongly acidic 10 ACS Paragon Plus Environment

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and basic conditions were detrimental to PMS-I– oxidation. To evaluate the performance under

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circumneutral conditions, we also examined the degradation of phenol in PMS solutions in which

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the pH values were further adjusted back to the specified values, accordingly, after adding PMS

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(Figure 2f). Consistent with the results in Figure 2e, strongly basic and acidic conditions were

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detrimental to PMS-I– oxidation: the degradation results at pH 3.8 and 10 were worse than those

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at pH 4.0 and 9.5, respectively. Lente et al.24 examined the oxidation kinetics of I– in PMS

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solution and found that the rate constant dropped sharply when the solution pH value was higher

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than 8.0. They related this pH dependence to the acid dissociation of PMS, which has a second

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ionization constant of 9.3 or 9.4;25, 26 in the solution pH range of 4 to 8, PMS mainly existed as

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its mononegative form (HSO ), whereas when the pH was higher than 11, almost all of the PMS

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existed in the form of a dianion (SO ). The deprotonated PMS (SO ) has a much lower

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3 –1 –1 24 reactivity ((3.0 ± 0.2) × 102 M–1 s–1) toward I– than HSO ((1.41 ± 0.03) × 10 M s ). In

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addition, compared with the performance around pH 4.5 and 4.9, the degradation of phenol at pH

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6 and 8 was markedly greater (Figure 2f). Therefore, it can be concluded that PMS-I– achieved

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its optimum degradation capability in the pH range from 6.0 to 8.0, representing a major

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advantage over the classical Fenton reagents, which only work effectively under strongly acidic

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conditions (around pH 3.5). Finally, we noted that the overall degradation of phenol after 10 min

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was only slightly influenced by the solution pH value within the range of 4 to 8.0 (Figures 2e and

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2f), which suggests that PMS-I– oxidation had a wide effective pH range.

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Figure 2. Effect of I– doses on phenol concentration (a) and its degradation kinetics (b) in PMS-I–

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oxidation; effect of I– doses on 2,4-diclorophenol concentration (c) and its degradation kinetics

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(d) in PMS-I– oxidation; effect of initial solution pH values without (e) and with (f) further pH

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adjustment after adding PMS. Conditions: [phenol] = 25 µM, [2,4-diclorophenol] = 12 µM,

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[PMS] = 65 µM, [Na2SO4] = 10 mM, and initial solution pH = 6.0 (a, b, c, and d; without further

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adjustment after adding PMS), and [I–] = 20 µM (e and f).


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Identification of Active Species and Degradation Mechanism. Analogously to the classical

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Fenton reactions, PMS has been widely used as a one-electron oxidant to generate ∙ in the

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presence of a reducing agent such as Fe2+ or Co2+.9 I– is also a reducing species and can thus

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activate PMS to generate oxygen-containing radicals. In addition, due to the asymmetrical

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structure of PMS, the simultaneous generation of both ∙ and ·OH via PMS decomposition is

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possible.27 Alternatively, ·OH can be produced by the reactions of ∙ with hydroxide (7.3 ×

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107 M–1 s–1)28 and water (103 to 104 s–1).15,

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were generated by PMS-I– oxidation and that these oxygen-containing radicals contributed to the

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degradation of the phenolic pollutants, although they should not be the dominant active species

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on the basis of the selective degradation results (Figure 1a). To evaluate the contribution of these

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radicals, methanol was used as a radical-scavenging probe, and its inhibitory effect on

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degradation was investigated. However, the results showed that methanol at concentrations up to

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2.5 M failed to inhibit the overall degradation of bisphenol A (Figure 3a). A similar failure of

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inhibition was also observed in the case of phenol degradation (Figure 3b). Conversely, phenol at

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a much lower concentration (1 to 5 mM) had a significant adverse effect on the degradation of

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bisphenol A (Figure 3c). Radicals such as hydroxymethyl radical (·CH2OH) and hydroxymethyl

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30 peroxy radical (·OOCH2OH) are probably formed during the oxidation of methanol by ∙ or

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·OH.31 The ·CH2OH reacts rapidly with oxygen to produce ·OOCH2OH that may undergo rapid

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self-reaction to generate formaldehyde and hydroperoxyl radical.31 To evaluate the potential role

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of these radicals in the degradation, we examine the effect of methanol on the degradation of

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9 phenol by PMS-Co2+ oxidation that is a well-known ∙ -dominated process. The results

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showed that over 90% of the phenol was degraded by PMS-Co2+ after 10 min, whereas no

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degradation was observed when methanol (2.5 M) presented (Figure S5). These results suggest

29

Therefore, it is possible that both ∙ and ·OH



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that phenol cannot be oxidized by the organic radicals produced from the oxidation of methanol

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6 –1 –1 30, 32 ∙ by ∙ and . Because methanol is relatively reactive toward ((3.2 to 9.7) × 10 M s )

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·OH (9.7 × 108 M–1 s−1),3 the failure of methanol to inhibit degradation suggests that neither of

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these radicals contributed to the degradation of the phenols. Formaldehyde, the oxidation product

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33 of methanol by ·OH or ∙ was also monitored. As displayed in Figure 3d, only ,

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approximately 0.35 µM formaldehyde was produced by PMS-I– oxidation with 325 µM PMS and

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100 µM I–. Considering the 90% degradation of 25 µM phenol by 65 µM PMS and 20 µM I–

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within 10 min (Figure 2a) and the extremely low concentration of formaldehyde produced with

283

– such high levels of PMS and I–, the contribution of ·OH or ∙ generated by PMS-I oxidation

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(if any) should be negligible.

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Quinones can be formed as degradation products of phenols and these compounds may

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activate PMS to generate 1O2.23 Although this active species might be produced in the PMS-I–

287

oxidation, it should not be the major oxidants due to the following reasons: (1) 1O2 reacts with I–

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at a rate constant of 8.7 × 105 M–1 s–1,34 and this rate is even lower than the reaction constant of

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1

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of 1O2 could not explain the inhibition effect of the slightly excess iodide ions. (2) According to

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Tratnyek and Holgne,35 chlorine-substituted phenols (e.g., 2-chlorophenol and 2,4-

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dichlorophenol) are more reactive toward 1O2, comparing to phenol. However, in PMS-I–

293

oxidation, the degradation of phenol (kapp = 1.01 min–1) was much faster than 2,4-dichlorophenol

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(kapp = 0.22 min–1), suggesting that 1O2 was not the major active oxidant. (3) Sodium azide is a

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specific quencher of 1O2 (4.5 × 108 M–1 s–1).36 This constant is over 100 times faster than that

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between 1O2 and phenol. The investigation on the effect of sodium azide showed that the

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presence of 10 µM sodium azide ([N3–]/[phenol] = 1:1) had no observable inhibition effect on the

O2 with 2,4-diclorophenol (5 × 106 M–1 s–1) or with phenol (3 × 106 M–1 s–1).35 The dominance


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degradation of phenol by PMS-I– oxidation (Figure S6). The result suggests that 1O2 should not

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be responsible for the degradation of phenol.

300 301

Figure 3. Effect of methanol on bisphenol A (a) and phenol (b) degradation in PMS-I– oxidation,

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effect of phenol on bisphenol A (c) degradation in PMS-I– oxidation, and production of

303

formaldehyde with methanol as the substrate in PMS-I– oxidation (d). Conditions: [bisphenol A]

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= 25 µM, [phenol] = 20 µM, [PMS] = 65 µM (a, b, and c), [I–] = 50 µM, [methanol] = 1 M (d), t

305

= 10 min (d), and initial solution pH = 6.0.

306 307

I– ions are known to be oxidized by ozone and chlorine to produce hypoiodous acid (HOI).37

308

Considering the strongly oxidative capability of PMS, HOI might be formed in PMS-I– oxidation 15 ACS Paragon Plus Environment


Page 16 of 29

309

reaction (Eq. 1). However, HOI is only capable of oxidizing phenol to iodophenols,38 which

310

cannot explain the rapid TOC removal from the phenol solution (Figure 1b). NH , a potential

311

scavenger of HOI, was further studied to evaluate the contribution of HOI, and the result showed

312

no influence of NH (up to 10 mM; [NH ]/[phenol] = 400) on the degradation of phenol (Figure

313

S7). These outcomes suggest that the other active species, instead of HOI, existed in the PMS-I–

314

oxidation.

315

Studies on the interactions between H2O2 and I– have shown that ·OH and I· are produced

316

via one-electron transfer (Eq. 2).39-41 Meanwhile, it is well established in the literature that PMS

317

is more easily activated than H2O2.42, 43 Therefore, a similar electron-transfer reaction generating

318

I· probably occurred during the interaction between PMS and I– (Eqs. 3 and 4). SO∙ was

319

generated during the activation of PMS by I–, but was rapidly scavenged by I– to produce I·. The

320

involvement of SO∙ could explain the slight inhibition effect of methanol (Figures. 3a and 3b).

321

Note that although the oxidation of I– by PMS has been investigated by several researchers over

322

the past decades,24, 44 none of them studied the fate of an organic compound in PMS-I– oxidation.

323

Because I· reacts very rapidly with compounds containing electron-rich groups, such as N-

324

methylindole ((1.9 ± 0.5) × 1010 M–1 s–1) and 4-methoxyphenol ((5.7 ± 1.2) × 109 M–1 s–1), via

325

one-electron transfer,15, 21 the pollutants containing a phenolic hydroxyl group(s), such as phenol,

326

were rapidly degraded under the attack of I· (Eq. 5). However, in addition to reacting with

327

phenolic compounds, the I· produced could also be rapidly scavenged by excess I– (Eq. 6) to

328

produce I∙ at a rate constant of 8.8 × 109 M–1 s–1,45 which was over 6 × 106 times faster than the

329

reaction between PMS and I– ((1.41 ± 0.03) × 103 M–1 s–1).24 The considerably faster reaction of

330

I· with I– than with PMS readily explains the dependence of the degradation rates on the

331

[PMS]/[I–] ratio, as observed in Figures 2a to 2d. Compared with the standard reduction potential 16 ACS Paragon Plus Environment

Page 17 of 29


332

of I ∙ ( E I∙ ⁄I = 1.33 V ),18 the potential of I∙ ( E I∙ ⁄I = 1.03 V )18 is much lower,

333

suggesting that the oxidative capability of I∙ is weaker than that of I·; an earlier pulse radiolysis

334

study showed that the reactivity between I∙ and natural phenolic compounds was too low to be

335

determined.21 More importantly, as shown in Eq. 7, I∙ underwent rapid self-quenching to

336

produce non-oxidative species such as I and I– ions at a rate constant of 2.3 × 109 M–1 s–1.41

337

Consequently, a slight overdose of I– (with [PMS]/[I–] decreased from 1.00 to 0.65) strongly

338

inhibited the degradation of phenols (Figures 2b and 2d) and a significant excess of I– ([I–]/[PMS]

339

= 154) near completely inhibited the degradation (Figure S8). Because the reactivity of RHS is

340

positively related to the electron-donating property of the target pollutant,21 the dominant role of

341

I ∙ in PMS-I– oxidation readily explains the pseudo-second-order kinetic differences in the

342

degradation of the selected phenols (Figures 1a and S4); the Hammett constants of hydroquinone,

343

phenol, and 2,4-dichlorophenol are -0.36,46 0,47 and 0.95,47 respectively, suggesting that the

344

electron-donating capability of these compounds is in the order hydroquinone > phenol > 2,4-

345

dichlorophenol, consistently with the degradation performances (Figure 1a).

346

HSO + I → HOI + SO 1

347

H O + I + H → ∙ OH + I ∙ + H O

(2)

348

∙ ∙ HSO + I → SO + I + OH

(3)

349 350 351 352

∙ SO∙ + I → SO + I

(4)

I ∙ + C% H OH → C% H O∙ + I + H

(5)

I ∙ + I → I∙

(6)

I∙ + I∙ → I + I


(7)


Page 18 of 29

353 354

Figure 4. Effects of phenolic pollutants on I production in PMS-I– oxidation. Inset shows

355

production of I in PMS-I– oxidation in the presence of 50 µM atrazine or 50 µM

356

chloramphenicol. Conditions: [PMS] = 100 µM, [I–] = 500 µM, [Na2SO4] = 10 mM, and initial

357

solution pH = 6.0.

358 359

On the basis of the proposed mechanism, I∙ was reduced to I by excess I– in the absence of

360

phenolic pollutants, and the production of I was assumed to be inhibited when phenols were

361

present as scavengers of I ∙ (Eq. 5). To verify this hypothesis, we spectrometrically monitored the

362

production of I under the influence of different phenols at 352 nm (Figure 4).48, 49 Without

363

phenolic pollutants, I was produced with an absorbance of about 0.366. In the presence of

364

phenol, the production of I was increasingly inhibited with the increase of the phenol

365

concentration from 1 to 50 µM. Resorcinol (Table S2) reacts more rapidly with I∙ ((1.3 ± 0.4) ×

366

108 M−1 s−1)21 than does phenol (1.6 × 107 M−1 s−1), and was thus expected to have a much more

367

pronounced effect. As shown in Figure 4, no production of I occurred in the presence of 50 µM

368

resorcinol; the yellow color indicating the formation of I also almost disappeared with 50 µM 18 ACS Paragon Plus Environment

Page 19 of 29


369

resorcinol (Figure S9a). To further prove the production of I , starch was used as an indicator

370

and added to the PMS-I– system with and without resorcinol. The results (Figure S9b) showed

371

that when starch (1 g L–1) was added, the color of the solution without resorcinol immediately

372

changed to blue, while the color of the solution with resorcinol did not have such dramatic

373

change. This observation again suggests the existence of I in the oxidation system without

374

resorcinol. In contrast to the strong inhibitory effect of the phenolic pollutants, no obvious

375

difference in the production of I was observed when 50 µM of the non-phenolic pollutants

376

atrazine or chloramphenicol was present (inset in Figure 4). These observations strongly suggest

377

the selective interactions of I∙ with phenols. Note that at the current stage, we have only

378

demonstrated the capability of I ∙ atoms to selectively mineralize phenolic pollutants (Figure 1b),

379

but the mineralization pathways remain unknown and will be investigated in our future research.

380 381

Production of Iodinated Intermediates. As shown in Figure 1a, phenol (10 µM) could be

382

completely degraded by PMS-I– oxidation after reaction for 4 min in the presence of 65 µM and

383

50 µM I–. Therefore, we investigated the production of iodinated intermediates under such

384

conditions with GC-MS and UPLC-MS/MS. The results only observed 2-iodophenol, 3-

385

iodophenol, and 2,4,6-trichloropenol as the iodinated degradation intermediates of phenol. Due

386

to the extremely low level of 2-iodophenol (< 0.045 µM after extraction), we only quantified 3-

387

iodophenol and 2,4,6-triiodophenol (Figure S10). The results showed that the yield of total

388

iodophenols produced was less than 2.0%. Thus, the production of halogenated byproducts

389

should not be a serious concern when using PMS-I– oxidation for phenol degradation, although

390

strategies to further minimize the production of these iodinated products will be surely beneficial



Page 20 of 29

391

in the future. In addition, special attention should be paid on the potential of generating more

392

persistent products, such as iodinated trihalomethanes, during the degradation of iodophenols.

393 394

Degradation of Phenols by PMS in the Presence of Cl–, Br–, or Inorganic Carbon. The effect

395

of halide ions (X–), including Cl– and Br–, on AOPs, particularly on UV/H2O2 oxidation, has

396

received extensive attention because they are common constituents in various water metrics

397

(Table S1).16, 17 Generally, these halide ions seriously inhibit the degradation of pollutants that

398

contain only electron-poor groups; they serve as sinks for both ·OH (Eqs. 8 and 9)3 and SO∙ (Eq.

399

10),10, 14, 16 and the halogen radical anions (X ∙) produced are relatively less reactive.18 However,

400

for pollutants that contain electron-rich groups, halide ions exert marginally inhibitory or even

401

promoting effects due to the greater selectivity of X ∙ and X ∙ radicals.17 Similarly to halide ions,

402

∙ 15, carbonates also readily react with both ∙ OH and SO∙ to produce CO ,

403

selective and highly reactive toward anilines and phenols.50, 51 In this study, we also examined

404

the extent of production of RHS and CO∙ in PMS solution when Cl–, Br–, or CO ions were

405

present. High doses of Cl– and Br– were used due to their relatively low reactivity toward PMS.44

406

When excess halide ions were present, their corresponding X ∙ atoms were transformed to X ∙

407

radicals (Eqs. 11 and 12). phenol was again used as a model pollutant due to its relatively rapid

408

8 –1 –1 14, 52 reactions with Cl∙ Br∙ (5.3 × 106 M–1 s–1),52 and CO∙ ((2.2 ± 0.2) ((2.5 to 2.7) × 10 M s ),

409

× 107 M–1 s–1).15

16

which is very

410

X + ∙ OH ↔ XOH ∙

(8)

411

XOH ∙ + H → X ∙ + H O

(9)

412

∙ SO∙ + X → X + SO

(10)

413

XOH ∙ + X → X∙ + OH

(11)

414

X ∙ + X → X∙

(12) 20

ACS Paragon Plus Environment

Page 21 of 29


415

Low levels of Cl– (1 to 10 mM) had a slight promoting effect on the degradation of phenol

416

by PMS alone (Figure 5a). However, the degradation was markedly enhanced when [Cl–] was

417

increased to 50 mM. Similar phenomena were also observed by Wang et al. when examining the

418

effect of Cl– on the decoloration of Orange II in PMS solution.53 Although traces of Br– and I–

419

were present in the supplied sodium chloride (as specified in the Chemicals section), their

420

presence had a negligible influence on the degradation of phenol (Figure S11). Br– interacted

421

much more strongly with PMS than did Cl–. As shown in Figure 5b, ~50% of the phenol was

422

degraded by PMS-Br– oxidation with 1 mM Br– after 10 min. However, this oxidative

423

combination was still less effective than PMS-I–; approximately 90% degradation of phenol was

424

achieved within 6 min with 50 µM I– (Figure 2a), i.e., just 5% of the Br– dose (1 mM). These

425

observations were consistent with previously reported kinetic results24 showing that the reaction

426

constants of PMS with I– ((1.41 ± 0.03) × 103 M–1 s–1), Br– ((7.0 ± 0.1) × 10–1 M–1 s–1), and Cl–

427

((2.06 ± 0.03) × 10–3 M–1 s–1) differed greatly. Because PMS reacts with Br– over 300 times

428

faster than with Cl–, the degradation of phenol by PMS-Br– oxidation (Figure 5b) was much

429

greater than that by PMS-Cl– oxidation (Figure 5a), even though Br∙ is less oxidative than Cl∙

430

in terms of their standard reduction potentials (E Br∙ ⁄2Br = 1.63 V and E Cl∙ ⁄2Cl =

431

2.09 V).15 HCO was used as a representative form of inorganic carbon. Before adding PMS, the

432

pH value of the phenol solution containing HCO was adjusted to 6.0. However, no obvious

433

effect on the degradation was observed when HCO at concentrations ranging from 1 to 50 mM

434

was added to a PMS solution containing phenol (Figure 5c). Previous studies showed that HCO

435

∙ ∙ ∙ 16, CO reacted efficiently with Cl and Br to produce CO .

436

∙ transformation of HCO to CO by reaction with I· is unlikely due to the higher reduction

437

54 potential of CO∙ E CO∙ than that of I· E I∙ ⁄I = 1.33.18 To verify ⁄CO = 1.59 V)


17

However, a similar


Page 22 of 29

438

this hypothesis, we investigated the degradation of atrazine by PMS-I– oxidation in the presence

439

– of HCO (CO ). As HCO (CO ) may influence PMS-I oxidation by providing background

440

ionic strength, we thus did not choose phenol degradation for this study. The choice of atrazine

441

as the target pollutant was based on the fact that this compound reacts rapidly with CO∙ ((3.7 to

442

4.0) × 106 M–1 s–1)51, 55 but cannot be degraded by I· (Figure 1a). As expected (Figure S12), no

443

degradation of atrazine by PMS-I– oxidation occurred when 10 to 50 mM bicarbonate was added,

444

confirming that CO∙ could not be generated by PMS-I–-HCO (CO ) oxidation.

445 –

446

Figure 5. Effects of Cl– (a), Br– (b), and HCO (c) on phenol degradation in PMS solutions.

447

Conditions: [PMS] = 65 µM, [phenol] = 25 µM, [Na2SO4] = 10 mM, and initial solution pH =

448

6.0.

449 450 451

IMPLICATIONS

452

I· is usually produced by laser flash photolysis. In this work, we demonstrated for the first time

453

the production of this RHS using the environmentally friendly and inexpensive PMS and the

454

readily available I–. The high reactivity and mineralizing capability of I· render PMS-I– oxidation

455

a promising AOP for the highly selective and metal-free removal of phenolic pollutants. Halide

456

ions are ubiquitous in various water matrices, particularly in hydraulic fracturing wastewater

457

(Table S1), in which extremely high levels of Cl–, Br–, and I– are co-present with phenolic


Page 23 of 29


458

pollutants.56 The generation of RHS by the simple addition of PMS into this wastewater may

459

represent a cost-effective and efficient decontamination method. However, if the high sulfate

460

concentration in the effluent may be a concern, an attention should be paid in using PMS-I–

461

oxidation for such remediation strategy, particularly when high levels of I– ions are present. The

462

spectrophotometric protocol based on I– oxidation has been widely used for the determination of

463

persulfates, including PMS and PDS, by quantifying the produced I . However, the results of this

464

study show that the production of I can be strongly inhibited by the presence of phenols, which

465

has usually been neglected in past research. Thus, care should be taken in using this approach to

466

quantify persulfate when high levels of phenols are present. PMS-I– can rapidly degrade phenolic

467

pollutants, which contain readily degradable phenolic hydroxyl moieties; anilines also bear

468

readily oxidizable amino group(s), and their degradation by PMS-I– oxidation should be

469

evaluated in future studies.

–

–

470

471

ASSOCIATED CONTENT

472

Supporting Information

473

Text S1–S4: Details regarding the setting of background ionic strength, analysis of the selected

474

pollutants, and identification of degradation intermediates and radicals; Table S1–S5: Halide ions

475

in various water sources, physicochemical properties of the selected pollutants, rate constants of

476

sulfite

477

pollutant/iodophenol analysis; Figure S1–S12: Effect of phosphate buffer, pollutant degradation

478

by PMS alone, selective degradation of phenol, pseudo-first-order constants, effect of methanol

479

on PMS-Co2+ oxidation, effects of sodium azide and NH4+ on phenol degradation, effect of

oxidation

by

oxidative

species,

UPLC-PDA/UPLC-MS/MS


conditions

for


480

resorcinol on I formation, profile of major iodinated intermediates, effect of Br– and I– in the

481

supplied sodium chloride on phenol degradation, and atrazine degradation by PMS-I–-HCO

482

(CO ) oxidation. This material is available free of charge on the ACS Publications website at

483

http://pub.acs.org.

484 485

AUTHOR INFORMATION

486

Corresponding Author

487

*Phone: +852-2859-1973; fax: +852-2559-5337; e-mail: [email protected] (K. Shih)

488

Notes

489

The authors declare no competing financial interest.

490

491

ACKNOWLEDGEMENTS

492

This work was supported by the General Research Fund (17206714, 17212015) and

493

Collaborative Research Fund (C7044-14G) of the Research Grants Council of Hong Kong. The

494

authors would like to thank Vicky Fung for her technical assistance and the anonymous

495

reviewers for their suggestions to improve the quality of this study.

496 497


Page 24 of 29

Page 25 of 29


498

REFERENCES

499 500 501 502 503 504 505 506 507 508 509 510 511 512 513 514 515 516 517 518 519 520 521 522 523 524 525 526 527 528 529 530 531 532 533 534 535 536 537 538 539 540 541 542

(1) Neyens, E.; Baeyens, J., A review of classic Fenton's peroxidation as an advanced oxidation technique. J. Hazard. Mater. 2003, 98 (1-3), 33-50. (2) Pignatello, J. J.; Oliveros, E.; MacKay, A., Advanced oxidation processes for organic contaminant destruction based on the Fenton reaction and related chemistry. Crit. Rev. Environ. Sci. Technol. 2006, 36 (1), 1-84. (3) Buxton, G. V.; Greenstock, C. L.; Helman, W. P.; Ross, A. B., Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals (⋅OH/⋅O−) in aqueous solution. J. Phys. Chem. Ref. Data 1988, 17 (2), 513-886. (4) Zhou, D.; Zhang, H.; Chen, L., Sulfur-replaced Fenton systems: Can sulfate radical substitute hydroxyl radical for advanced oxidation technologies? J. Chem. Technol. Biotechnol. 2015, 90 (5), 775-779. (5) Haag, W. R.; Yao, C. C. D., Rate constants for reaction of hydroxyl radicals with several drinking water contaminants. Environ. Sci. Technol. 1992, 26 (5), 1005-1013. (6) Wang, G. S.; Hsieh, S. T.; Hong, C. S., Destruction of humic acid in water by UV light— catalyzed oxidation with hydrogen peroxide. Water Res. 2000, 34 (15), 3882-3887. (7) Gultekin, I.; Ince, N. H., Degradation of reactive azo dyes by UV/H2O2: Impact of radical scavengers. J. Environ. Sci. Health., Part A 2004, 39 (4), 1069-1081. (8) Autin, O.; Hart, J.; Jarvis, P.; MacAdam, J.; Parsons, S. A.; Jefferson, B., The impact of background organic matter and alkalinity on the degradation of the pesticide metaldehyde by two advanced oxidation processes: UV/H2O2 and UV/TiO2. Water Res. 2013, 47 (6), 2041-2049. (9) Anipsitakis, G. P.; Dionysiou, D. D., Degradation of organic contaminants in water with sulfate radicals generated by the conjunction of peroxymonosulfate with cobalt. Environ. Sci. Technol. 2003, 37 (20), 4790-4797. (10) Feng, Y.; Wu, D.; Deng, Y.; Zhang, T.; Shih, K., Sulfate radical-mediated degradation of sulfadiazine by CuFeO2 rhombohedral crystal-catalyzed peroxymonosulfate: Synergistic effects and mechanisms. Environ. Sci. Technol. 2016, 50 (6), 3119-27. (11) Oh, W. D.; Dong, Z. L.; Lim, T. T., Generation of sulfate radical through heterogeneous catalysis for organic contaminants removal: Current development, challenges and prospects. Appl. Catal., B 2016, 194, 169-201. (12) Zhang, T.; Chen, Y.; Wang, Y.; Le Roux, J.; Yang, Y.; Croué, J. P., Efficient peroxydisulfate activation process not relying on sulfate radical generation for water pollutant degradation. Environ. Sci. Technol. 2014, 48 (10), 5868-5875. (13) Duan, X.; Sun, H.; Wang, Y.; Kang, J.; Wang, S., N-doping-induced nonradical reaction on single-walled carbon nanotubes for catalytic phenol oxidation. ACS Catal. 2014, 5, (2), 553-559. (14) Hasegawa, K.; Neta, P., Rate constants and mechanisms of reaction of chloride (Cl2-) radicals. J. Phys. Chem. 1978, 82 (8), 854-857. (15) Neta, P.; Huie, R. E.; Ross, A. B., Rate constants for reactions of inorganic radicals in aqueous solution. J. Phys. Chem. Ref. Data 1988, 17 (3), 1027-1284. (16) Yang, Y.; Pignatello, J. J.; Ma, J.; Mitch, W. A., Comparison of halide impacts on the efficiency of contaminant degradation by sulfate and hydroxyl radical-based advanced oxidation processes (AOPs). Environ. Sci. Technol. 2014, 48 (4), 2344-2351. (17) Grebel, J. E.; Pignatello, J. J.; Mitch, W. A., Effect of halide ions and carbonates on organic contaminant degradation by hydroxyl radical-based advanced oxidation processes in saline waters. Environ. Sci. Technol. 2010, 44 (17), 6822-6828. 25 ACS Paragon Plus Environment


543 544 545 546 547 548 549 550 551 552 553 554 555 556 557 558 559 560 561 562 563 564 565 566 567 568 569 570 571 572 573 574 575 576 577 578 579 580 581 582 583 584 585 586 587

Page 26 of 29

(18) Stanbury, D. M., Reduction potentials involving inorganic free radicals in aqueous solution. Adv. Inorg. Chem. 1989, 33, 69-138. (19) Langmuir, M. E.; Hayon, E., Flash photolysis study of mercury (II) halide complexes in aqueous solutions. Rates of reaction X2- radical antons. J. Phys. Chem. 1967, 71, (12), 3808-3814. (20) Von Gunten, U., Ozonation of drinking water: Part II. Disinfection and by-product formation in presence of bromide, iodide or chlorine. Water Res. 2003, 37 (7), 1469-1487. (21) Alfassi, Z.; Huie, R.; Marguet, S.; Natarajan, E.; Neta, P., Rate constants for reactions of iodine atoms in solution. Int. J. Chem. Kinet. 1995, 27 (2), 181-188. (22) Zhou, X.; Mopper, K., Determination of photochemically produced hydroxyl radicals in seawater and freshwater. Mar. Chem. 1990, 30, 71-88. (23) Zhou, Y.; Jiang, J.; Gao, Y.; Ma, J.; Pang, S. Y.; Li, J.; Lu, X. T.; Yuan, L. P., Activation of peroxymonosulfate by benzoquinone: A novel nonradical oxidation process. Environ. Sci. Technol. 2015, 49 (21), 12941-12950. (24) Lente, G.; Kalmár, J.; Baranyai, Z.; Kun, A.; Kék, I.; Bajusz, D.; Takács, M.; Veres, L.; Fábián, I., One- versus two-electron oxidation with peroxomonosulfate ion: Reactions with iron(II), vanadium(IV), halide ions, and photoreaction with cerium(III). Inorg. Chem. 2009, 48 (4), 1763-1773. (25) Goodman, J. F.; Robson, P., Decomposition of inorganic peroxyacids in aqueous alkali. J. Chem. Soc. (Resumed) 1963, (0), 2871-2875. (26) Ball, D. L.; Edwards, J. O., The kinetics and mechanism of the decomposition of Caro's acid. I. J. Am. Chem. Soc. 1956, 78 (6), 1125-1129. (27) Kennedy, R. J.; Stock, A. M., The oxidation of organic substances by potassium peroxymonosulfate. J. Org. Chem. 1960, 25 (11), 1901-1906. (28) Chawla, O. P.; Fessenden, R. W., Electron spin resonance and pulse radiolysis studies of some reactions of peroxysulfate (SO4.-). J. Phys. Chem. 1975, 79 (24), 2693-2700. (29) Liang, C.; Su, H. W., Identification of sulfate and hydroxyl radicals in thermally activated persulfate. Ind. Eng. Chem. Res. 2009, 48 (11), 5558-5562. (30) Clifton, C. L.; Huie, R. E., Rate constants for hydrogen abstraction reactions of the sulfate radical, SO4-. alcohols. Int. J. Chem. Kinet. 1989, 21 (8), 677-687. (31) Monod, A.; Chebbi, A.; Durand-Jolibois, R.; Carlier, P., Oxidation of methanol by hydroxyl radicals in aqueous solution under simulated cloud droplet conditions. Atmos. Environ. 2000, 34 (29–30), 5283-5294. (32) Eibenberger, H.; Steenken, S.; O'Neill, P.; Schulte-Frohlinde, D., Pulse radiolysis and electron spin resonance studies concerning the reaction of SO4. with alcohols and ethers in aqueous solution. J. Phys. Chem. 1978, 82 (6), 749-750. (33) Keenan, C. R.; Sedlak, D. L., Factors affecting the yield of oxidants from the reaction of nanoparticulatezero-valent iron and oxygen. Environ. Sci. Technol. 2008, 42 (4), 1262-1267. (34) Braathen, G.; Chou, P. T.; Frei, H., Time-resolved reaction of oxygen (1. DELTA.) with iodide in aqueous solution. J. Phys. Chem. 1988, 92 (23), 6610-6615. (35) Tratnyek, P. G.; Hoigne, J., Oxidation of substituted phenols in the environment: A QSAR analysis of rate constants for reaction with singlet oxygen. Environ. Sci. Technol. 1991, 25 (9), 1596-1604. (36) Miskoski, S.; García, N. A., Influence of the peptide bond on the singlet molecular oxygenmediated (O2 [g]) photooxidation of histidine and methionine dipeptides. A kinetic study. Photochem. Photobiol. 1993, 57 (3), 447-452.


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(37) Bichsel, Y.; Von Gunten, U., Oxidation of iodide and hypoiodous acid in the disinfection of natural waters. Environ. Sci. Technol. 1999, 33 (22), 4040-4045. (38) Dietrich, A. M.; Mirlohi, S.; DaCosta, W. F.; Dodd, J. P.; Sauer, R.; Homan, M.; Schultz, J., Flavor profile analysis and GC/MS detection of phenolic iodinated disenfection byproducts in drinking water for the USA space program. Water Sci. Technol. 1999, 40 (6), 45-51. (39) Stanisavljev, D. R.; Milenkovic, M. C.; Mojović, M. D.; Popović-Bijelić, A. D., Oxygen centered radicals in iodine chemical oscillators. J. Phys. Chem. A 2011, 115 (27), 7955-7958. (40) Stanisavljev, D. R.; Milenkovic, M. C.; Mojović, M. D.; Popović-Bijelić, A. D., A potential source of free radicals in iodine-based chemical oscillators. J. Phys. Chem. A 2011, 115 (11), 2247-2249. (41) Milenković, M. C.; Stanisavljev, D. R., Role of free radicals in modeling the iodideperoxide reaction mechanism. J. Phys. Chem. A 2012, 116 (23), 5541-5548. (42) Anipsitakis, G. P.; Dionysiou, D. D., Radical generation by the interaction of transition metals with common oxidants. Environ. Sci. Technol. 2004, 38 (13), 3705-3712. (43) Yang, S.; Wang, P.; Yang, X.; Shan, L.; Zhang, W.; Shao, X.; Niu, R., Degradation efficiencies of azo dye Acid Orange 7 by the interaction of heat, UV and anions with common oxidants: Persulfate, peroxymonosulfate and hydrogen peroxide. J. Hazard. Mater. 2010, 179 (1–3), 552-558. (44) Fortnum, D. H.; Battaglia, C. J.; Cohen, S. R.; Edwards, J. O., The kinetics of the oxidation of halide ions by monosubstituted peroxides. J. Am. Chem. Soc. 1960, 82 (4), 778-782. (45) Elliot, A. J., A pulse radiolysis study of the reaction of OH with I2 and the decay of I2. Can. J. Chem. 1992, 70 (6), 1658-1661. (46) Schwarzenbach, R. P.; Gschwend, P. M.; Imboden, D. M., Environmental organic chemistry. John Wiley & Sons: 2005. (47) Adams, C. D.; Cozzens, R. A.; Kim, B. J., Effects of ozonation on the biodegradability of substituted phenols. Water Res. 1997, 31 (10), 2655-2663. (48) Liang, C.; Huang, C.; Mohanty, N.; Kurakalva, R. M., A rapid spectrophotometric determination of persulfate anion in ISCO. Chemosphere 2008, 73 (9), 1540-1543. (49) Milenkovic, M. C.; Potkonjak, N. I., The effect of hydroxycinnamic acids on oxy-radical generating iodide–hydrogen peroxide reaction. Bull. Chem. Soc. Jpn. 2014, 87 (11), 1255-1259. (50) Chen, S. N.; Hoffman, M. Z.; Parsons Jr, G. H., Reactivity of the carbonate radical toward aromatic compounds in aqueous solution. J. Phys. Chem. 1975, 79 (18), 1911-1912. (51) Canonica, S.; Kohn, T.; Mac, M.; Real, F. J.; Wirz, J.; von Gunten, U., Photosensitizer method to determine rate constants for the reaction of carbonate radical with organic compounds. Environ. Sci. Technol. 2005, 39 (23), 9182-9188. (52) Alfassi, Z.; Huie, R.; Neta, P.; Shoute, L., Temperature dependence of the rate constants for reaction of inorganic radicals with organic reductants. J. Phys. Chem. 1990, 94 (25), 8800-8805. (53) Wang, Z.; Yuan, R.; Guo, Y.; Xu, L.; Liu, J., Effects of chloride ions on bleaching of azo dyes by Co2+/oxone regent: Kinetic analysis. J. Hazard. Mater. 2011, 190 (1), 1083-1087. (54) Huie, R. E.; Clifton, C. L.; Neta, P., Electron transfer reaction rates and equilibria of the carbonate and sulfate radical anions. Int. J. Radiat. Appl. Instrum. C Radiat. Phys. Chem. 1991, 38 (5), 477-481. (55) Huang, J.; Mabury, S. A., A new method for measuring carbonate radical reactivity toward pesticides. Environ. Toxicol. Chem. 2000, 19 (6), 1501-1507.



632 633 634

(56) Lester, Y.; Ferrer, I.; Thurman, E. M.; Sitterley, K. A.; Korak, J. A.; Aiken, G.; Linden, K. G., Characterization of hydraulic fracturing flowback water in Colorado: Implications for water treatment. Sci. Total Environ. 2015, 512–513, 637-644.

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