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Hydroxylamine Promoted Goethite Surface Fenton Degradation of Organic Pollutants Xiaojing Hou, Xiaopeng Huang, Falong Jia, Zhihui Ai, Jincai Zhao, and Lizhi Zhang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b05906 • Publication Date (Web): 30 Mar 2017 Downloaded from http://pubs.acs.org on March 31, 2017

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Environmental Science & Technology

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Hydroxylamine Promoted Goethite Surface Fenton

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Degradation of Organic Pollutants

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Xiaojing Hou, Xiaopeng Huang, Falong Jia, Zhihui Ai, Jincai Zhao, and Lizhi Zhang*

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Key Laboratory of Pesticide & Chemical Biology of Ministry of Education, Institute of

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Environmental & Applied Chemistry, Central China Normal University, Wuhan 430079, P. R. China

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RECEIVED DATE (to be automatically inserted after your manuscript is accepted if required according to the journal that you are submitting your paper to)

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* To whom correspondence should be addressed. E-mail: [email protected]. Phone/Fax: +86-27-6786 7535. 1

ACS Paragon Plus Environment

Environmental Science & Technology 24

ABSTRACT: In this study, we construct a surface Fenton system with hydroxylamine (NH2OH),

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goethite (α-FeOOH), and H2O2 (α-FeOOH-HA/H2O2) to degrade various organic pollutants

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including dyes (methyl orange, methylene blue, and rhodamine B), pesticides (pentachlorophenol,

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alachlor, and atrazine), and antibiotics (tetracycline, chloramphenicol, and lincomycin) at pH 5.0. In

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this surface Fenton system, the presence of NH2OH could greatly promote the H2O2 decomposition

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on the α-FeOOH surface to produce •OH without releasing any detectable iron ions during the

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alachlor degradation, which was different from some previously reported heterogeneous Fenton

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counterparts. Moreover, the •OH generation rate constant of this surface Fenton system was 102 −

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104 times those of previous heterogeneous Fenton processes. The interaction between α-FeOOH and

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NH2OH was investigated with using attenuated total reflectance Fourier transform infrared

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spectroscopy and density functional theory calculations. The effective degradation of organic

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pollutants in this surface Fenton system was ascribed to the efficient Fe(III)/Fe(II) cycle on the

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α-FeOOH surface promoted by NH2OH, which was confirmed by X-ray photoelectron spectroscopy

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analysis. The degradation intermediates and mineralization of alachlor in this surface Fenton system

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were then systematically investigated using total organic carbon and ion chromatography, liquid

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chromatography-mass spectrometry and gas chromatography-mass spectrometry. This study offers a

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new strategy to degrade organic pollutants, and also sheds light on the environmental effects of

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goethite.

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Key Words: Surface Fenton; Goethite; Hydroxylamine; Alachlor; Degradation.

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INTRODUCTION

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To avoid the drawbacks of traditional homogenous Fenton (Fe2+/H2O2) systems, such as the ferric

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oxide sludge generation, the strait working pH value of 2.0 – 3.5 and the damage to the catalyst, 2


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heterogeneous Fenton systems are designed with iron bearing catalysts for pollutants control and

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environmental remediation.1-4 However, the Fe(III)/Fe(II) cycle of most iron bearing catalysts during

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heterogeneous Fenton process is not as effective as that of traditional homogeneous Fenton

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counterpart, lowering the efficiencies of H2O2 decomposition and hydroxyl radicals (•OH)

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generation, and thus restricting the application of heterogeneous Fenton process. For instance, the

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•OH generation rate constants of heterogeneous Fenton reagents such as goethite, hematite and

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ferrihydrite were as low as 4.00 × 10-7, 4.25 × 10-5, and 2.00 × 10-5 s-1, respectively.5

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In order to promote the efficiency of heterogeneous Fenton process, chelating or reducing agents

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were often employed to enhance the Fe(III)/Fe(II) cycle of iron oxide in heterogeneous Fenton

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reactions. For instance, the addition of carboxy-methyl-β-cyclodextrin (CMCD) could improve the

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degradation of 2,4,6-trinitrophenolin by a factor of 3 in the magnetite-bearing mineral heterogeneous

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Fenton system, because CMCD could favor the dissolution of iron ions from mineral surface.6

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Moreover, Xue et al. reported that the pentachlorophenol degradation rate in the magnetite

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heterogeneous Fenton process increased by 1.7 – 5.7 times after adding chelating agents such as

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succinate, citrate, tartrate, CMCD, ethylenediaminetetraacetic acid (EDTA), and oxalate.7 They

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attributed the enhanced efficiency of heterogeneous Fenton system to iron ions absorbed on the

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catalyst surface, rather than more dissolved iron ions in the solution. For example, EDTA and

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CMCD could more strongly complex with iron ions than oxalate, and thus promote the dissolution

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of iron bearing minerals, increasing the dissolved iron ions concentrations to the range of 2.8 – 14

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mg/L during the heterogeneous Fenton processes, but oxalate was found to be more efficient to

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improve the efficiency of heterogeneous Fenton system than EDTA and CMCD. Interestingly, Wang

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et al. found that mesoporous copper ferrite could effectively catalyze the H2O2 decomposition

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without significantly releasing iron ions, as the concentration of iron ions during the heterogeneous

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Fenton processes was less than 1 mg/L even under acidic condition. More importantly, this catalyst 3



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retained its high catalytic activity even after 5 cycles of use.8 According to the concentration of iron

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ions released during the heterogeneous Fenton processes, herein we intend to classify heterogeneous

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Fenton systems into heterogeneous Fenton-like and surface Fenton ones. In the heterogeneous

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Fenton-like systems of considerable dissolved iron ions in the solution, the Fe(III)/Fe(II) cycle, the

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generation of reactive oxygen species (ROS), and the pollutant oxidation might be similar with those

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in homogeneous Fenton systems.6 However, in the surface Fenton systems with limited dissolved

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iron ions (< 1 mg/L), the surface Fe(III)/Fe(II) cycle is crucial for the H2O2 decomposition and the

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pollutant degradation. Many studies reported that surface Fe(II) bound on iron oxides exhibited

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higher reactivity than dissolved Fe(II). For example, Klausen et al. demonstrated that the reduction

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of substituted nitrobenzenes could only occur by the reduction of Fe(II) adsorbed on iron

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(hydr)oxide surfaces, rather than the dissolved ferrous ions.9 Hofstetter et al. also found that the

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mineral-bound Fe(II) species could reduce the nitroaromatic compounds efficiently.10 Recently, our

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group reported that hematite facet confined Fe(II) was more active than dissolved Fe(II) for peroxide

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conversion, because ferrous ions confined on the hematite facets could significantly promote the

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H2O2 decomposition to produce •OH by lowering H2O2 decomposition energetic span.11 Although

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the surface Fenton systems could save iron resources and extend working pH range as well as avoid

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excess reductant iron dissolution, to achieve higher ROS utilization efficiency and better water safety,

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we still need to clarify the mechanisms of surface Fenton systems, and thus deeply understand the

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related Fe(III)/Fe(II) cycle, ROS generation, and pollutant oxidation pathways.

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Hydroxylamine (HA) is a common reductive chemical and also a kind of hydroxamates as an

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essential growth factor of some microbes.12, 13 Recently, HA was used to enhance the Fe(III)/Fe(II)

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cycle of Fe(III)/H2O2 and Fe(II)/PMS systems for the oxidative benzoic acid degradation.14,

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Meanwhile, HA of high concentration (10 mmol/L) alone could decompose H2O2 to produce •OH

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because of its reducibility.16 Furthermore, HA was more efficient than other reductive agents such as 4


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ascorbic acid, oxalate, p-hydroquinone, humus acid, and gallic acid, to degrade 2, 4,

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6-tribromophennol in a heterogeneous Fenton system under pH 3,17 because HA did not consume

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•OH as much as other reductive agents. Despite these advances, the interaction between HA and iron

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containing minerals is still unknown, which hampers our understanding on the Fe(III)/Fe(II) cycle

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and the pollutant removal in the heterogeneous Fenton system composed of HA and iron containing

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minerals.

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Herein we utilize goethite, H2O2 and HA with relatively low concentration (0.5 mmol/L) to

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construct a highly efficient surface Fenton system, aiming to remove various organic pollutants

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including dyes (methyl orange, methylene blue, and rhodamine B), pesticides (pentachlorophenol,

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alachlor, and atrazine), and antibiotics (tetracycline, chloramphenicol, and lincomycin) at pH 5. The

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Fe(III)/Fe(II) cycle, the ROS generation, and the alachlor degradation pathway are investigated in

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detail. The stability and the reusability of goethite are also investigated.

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MATERIALS AND METHODS

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Chemicals. HCl, NaOH, H2O2 (30 wt%), rhodamine B, methyl orange, iso-propanol, 1,

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10-phenanthroline, and 2, 2’-bipyridine were of analytical grade and bought from Medicines

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Corporation Ltd. China National. Goethite (α-FeOOH) was bought from Alfa Aesar.

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Dichloromethane, acetone and acetonitrile were all of HPLC grade and purchased from Merck

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KGaA. Hydroxylamine hydrochloride (HA, 99.9%), alachlor, atrazine, pentachlorophenol,

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lincomycin, tetracycline and chloramphenicol were purchased from Sigma-Aldrich. 0.1 mol/L of

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H2O2 stock solution was prepared by diluting 30 wt% H2O2. 0.1 mol/L of HA stock solution was

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produced by dissolving HA in deionized and oxygen-free water. All the concentrations of organic

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pollutant solutions were 20 mg/L. The NaOH and HCl solutions were utilized to adjust the pH

5



values of degradation solutions. All the stock solutions were freshly prepared before use and covered

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with foil paper to avoid light.

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Degradation Procedures. Batch trials were carried out in 100 mL conical flasks under magnetic

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stirring at 25 ± 2 ℃. Briefly, the reactions were triggered by adding 5 mg of α-FeOOH, 250 µL of 0.1

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mol/L HA solution and 500 µL of 0.1 mol/L H2O2 solution into 50 mL of alachlor solution (20

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mg/L) in sequence. After that, 900 µL of the reaction solution was sampled at predetermined

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intervals, following with the filtration through 0.22 µm filter membranes. Subsequently, ethanol (100

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µL) was added immediately into the sampled solutions to stop the reaction for the alachlor

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concentration measurement. The initial pH value of α-FeOOH-HA/H2O2 system was 5.0 without

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adjusting. HCl and NaOH were used to adjust the initial pH values of the reaction solutions. The

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anaerobic experiments were conducted by bubbling Ar gas to remove dissolved oxygen. Briefly,

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conical flasks sealed with the rubber stoppers were pumped with high purity Ar gas, along with

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needles being inserted in the rubber stoppers for gas escaping. To assess the stability of α-FeOOH,

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the catalyst after reaction was separated from the solution, then washed with deionized water and

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ethanol thoroughly, finally vacuum dried for the reuse. Besides alachlor, the degradation

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experiments of other organic pollutants, including rhodamine B, methyl orange, atrazine,

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pentachlorophenol, lincomycin, tetracycline and chloramphenicol, were also conducted. All the other

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experimental procedures were the same as those of alachlor degradation.

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Analytical Methods. The analysis procedures of organic pollutants were described in the supporting

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information. The concentrations of total dissolved irons and surface Fe(II) were determined by a 1,

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10-phenanthroline method with a UV–vis spectrophotometer (UV-2550, Shimadzu, Japan) unless it

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was specially stated.18 Hydroxylamine hydrochloride was used for the measurement of total iron

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concentration. Briefly, 0.5 mL of the resulting solution was taken, following with filtration through

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0.22 µm filter membranes. 0.5 mL of hydroxylamine hydrochloride solution (100 g/L) was then 6


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added. After 5 min of reduction, 0.5 mL of 1,10-phenanthroline solution (2 g/L) and 0.5 mL of

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sodium acetate solution (10%, w:w) were added in sequence. After 30 min of chromogenic reaction,

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the total dissolved irons amount was then analyzed. For the dissolved Fe(II) detection, 0.5 mL of

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hydroxylamine hydrochloride solution (100 g/L) was replaced with 0.5 mL of H2O. Then,

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Fe(II)-1,10-phenanthroline complex was measured at λ = 510 nm by the UV–vis spectrophotometer

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with the detection limit of 0.47 µmol/L (0.026 mg/L). Surface Fe(II) was determined with a modified

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10-phenanthroline method after careful pretreatment.19,

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reaction with HA for 1 h was deaerated by bubbling Ar gas for 10 min before being transferred to

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the centrifuge tube in a glovebox under anaerobic condition. Subsequently, the supernatant was

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discarded and the solid sample in the centrifuge tube was rinsed with neutral oxygen-free water for

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twice. All the operations were done in the anaerobic glovebox except the centrifugation process. As

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HA are easily dissolved in water and iron ions on the α-FeOOH surface are not soluble in water

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under neutral condition, HA in the α-FeOOH slurry and on the α-FeOOH surface could be removed

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with this process, which was confirmed by measuring HA and iron ions in the supernatants (SI

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Figure S1). As shown in Figure S1, iron ions were not detected in the primary and secondary

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supernatants and HA was not detected in the secondary supernatant. Subsequently, the resulting

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α-FeOOH slurry was extracted with 1 mol/L HCl aqueous solution under magnetic stirring and the

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protection of Ar gas for 1 h. Then, the resulting supernatant was measured with the same 1,

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10-phenanthroline method described above for the accumulated surface Fe(II) content. The Fe(II)

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content (µmol/g) generated on the α-FeOOH surface in the presence of HA was calculated by the

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following equation.

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Typically, the α-FeOOH slurry after

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Fe(II)Surface (µmol/g) = [A (µmol/L) × V (L)]/Wα-FeOOH (g)

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Where A is the concentration of Fe(II) in the resulting solution after pretreatment (µmol/L), V is

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the volume of the resulting solution after pretreatment (L), Wα-FeOOH is the mass of α-FeOOH (g). To 7


(1)


avoid the possible Fe(III) photochemical reduction, all the samples were kept in dark thoroughly.

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H2O2 concentration was determined by a p-hydroxyphenyl acetic acid fluorescence method with

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using a Fluoro Max-P spectrophotometer (FL1008M018, Cary, USA).21 A spectrophotometric

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method was used for HA concentration measuration.22 DMPO–•OH electron spin resonance (ESR)

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spectra were obtained with a JES FA 200 X-band spectrometer (JEOL, Japan). Anions in the reaction

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solution were detected by an ion chromatograph (IC, Dionex ICS-900, Thermo) equipped with an

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AS23 column. The injection volume was 10 µL and the flow rate was 1.0 mL/min. The mobile phase

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was the solution of Na2CO3 (4.5 mmol/L) and NaHCO3 (0.8 mmol/L). The pH values were measured

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with a pH meter (Thermo Orion 720A+).

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The total organic carbon (TOC) content change of alachlor in the α-FeOOH-HA/H2O2 system was

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detected with a Shimadzu TOC-V CPH analyzer. Briefly, the degradation experiment of alachlor was

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conducted in a 100-mL conical flask under agitating at room temperature. 4 mL of the suspension

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was taken at predetermined time intervals, following with the filtration through 0.22 µm filter

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membranes. The Brunauer−Emmett−Teller (BET) surface area of α-FeOOH was calculated from N2

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adsorption/desorption isotherms with a specific surface area analyzer (Micromeritics Tristar 3000).

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The concentrations of dissolved iron ions were detected with inductively coupled plasma optical

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emission spectrometer (ICP-OES, Agilent, Varian 700) in some cases. The detection limit of

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ICP-OES was 0.035 µmol/L (0.002 mg/L). The element chemical states on α-FeOOH surface were

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examined by the X-ray photoelectron spectroscopy (XPS, PHI Quantera II), which was described in

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the supporting information.

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The solutions were sampled for the degradation intermediate analysis when alachlor was degraded

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in the α-FeOOH-HA/H2O2 system for 30 min. The possible alachlor degradation intermediates were

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identified by a gas chromatography-mass spectrometry (GC-MS, Trace 1300, ISQ, Thermo) with a

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DB-5 column (30 m × 0.25 mm) and a liquid chromatography-mass spectrometry (LC-MS, TSQ 8


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Quantum MAX, Thermo) with a Hypersil Gold C18 column (150 mm × 2.1 mm, 300 Å). The

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pre-treatment methods and the subsequent analysis process were reported previously.23

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Attenuated total reflectance Fourier transform infrared (ATR-FTIR) spectra were recorded using a

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FT-IR spectrometer (Nicolet iS50, Thermo) equipped with a diamond internal reflection element and

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a sensitive MCT detector cooled by liquid N2. α-FeOOH films were prepared by using a modified

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deposition method.24 Typically, 10 µL of α-FeOOH suspension (1 g/L) was dropped onto the

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diamond internal reflection element and dried to obtain semitransparent α-FeOOH films. A

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background spectrum was then collected. Subsequently, IR spectroscopy were obtained after the

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α-FeOOH film was reacted with HA at pH 5.0. HA-iron complex models were calculated using

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density functional theory (DFT) with Gaussian 09.25 The optimization and frequencies calculation of

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iron oxide cluster models were calculated with using the method described in our previous work.24

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RESULTS AND DISCUSSION

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The coordination pattern of HA on α-FeOOH surface was first investigated via ATR-FTIR

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spectroscopy to understand the HA absorption behavior on the α-FeOOH surface and their

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interaction. In the IR spectra of HA aqueous solution (SI Figure S2a), the peaks at 1102 and 1024

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cm-1 were generated from the N-H bending vibration, and the peaks at 906 cm-1 and 796 cm-1 were

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arisen from the N-O stretch and the O-H bending vibration, respectively.26 The stretching

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frequencies of these peaks did not shift along with increasing HA concentration, suggesting that they

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were less sensitive to the HA concentration. ATR-FTIR was then used to record the IR spectra of HA

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adsorbed on the α-FeOOH film (SI Figure S2b). In the IR spectra of HA adsorbed on α-FeOOH film,

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the N-H bending vibration peaks were located at 1137, 1044, and the N-O stretch and O-H bending

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vibration peaks appeared at 899 and 789 cm-1 (SI Figure S2c), respectively, which were slightly

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different from those of HA aqueous solutions. These differences indicated the formation of 9



inner-sphere iron-HA complexes on the α-FeOOH surface. Moreover, the intensities and the

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positions of four peaks did not change in the presence of NaCl (SI Figure S2d), further confirming

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that the inner iron-HA complexes were formed on the α-FeOOH surface. To further identify the

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specific coordination pattern of iron-HA complexes on the α-FeOOH surface, the vibrational

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frequencies of some possible iron-HA complexes models such as iron-OHNH2 and iron-NH2OH

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were calculated to fit for the HA binding models on α-FeOOH. We found that the calculated

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vibrational frequencies of iron-OHNH2 complexes matched better with the experimental ones than

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that of iron-NH2OH complexes (SI Figure S3 and Table S1), as the values of R2 and slope were

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closer to 1, and the values of intercept and standard deviation were smaller in case of iron-OHNH2

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complexes. Thus, we believed that the iron-OHNH2 complexes were formed on the α-FeOOH

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surface. By analyzing these data, we were able to clarify the specific absorption behavior of HA on

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the α-FeOOH surface and how the redox reaction happened between HA and Fe(III) on the

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α-FeOOH surface. As the model of HA adsorption on the α-FeOOH surface highly determined the

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kinetics of Fe(II) generation on the α-FeOOH surface, we could understand that the complex of

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≡Fe(III)-HA formed on the α-FeOOH surface was the first step of Fe(II) generation on the α-FeOOH

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surface, according to the adsorption model determined with FTIR spectra and DFT computation

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results.

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We subsequently investigated the α-FeOOH reduction process in the presence of HA by

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monitoring the chemical state changes of iron species in the aqueous solution and on the α-FeOOH

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surface. Although neither Fe(III) nor Fe(II) was released from α-FeOOH in the HA solution (Figure

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1a), the hydrochloric acid extraction experimental results revealed the existence of Fe(II) on the

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α-FeOOH surface (Figure 1b), suggesting that the surface Fe(III) of α-FeOOH was reduced by HA.

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Moreover, the presence of molecular oxygen could slightly affect the concentration of Fe(II)

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extracted from α-FeOOH pre-treated with HA solution of different concentrations, as surface Fe(II) 10


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could be oxidized by molecular oxygen. Therefore, the interaction between α-FeOOH and HA could

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be described as follows. First, inner-sphere iron-HA complexes (≡Fe(III)-HA) were formed on the

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α-FeOOH surface (Eq. 2). Then, surface Fe(II) (≡Fe(II)-OH2) was generated by the reduction of HA

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(Eq. 3).14, 27

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≡Fe(III)-OH2 + HA ≡Fe(III)-HA + H2O

(2)

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≡Fe(III)-HA + H2O ≡Fe(II)-OH2 + •HA

(3)

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Regarding that HA could induce the Fe(III) reduction on the α-FeOOH surface, we thus

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constructed a surface Fenton system with α-FeOOH, HA, and H2O2 (α-FeOOH-HA/H2O2) to

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degrade various organic pollutants including dyes (methyl orange, methylene blue, and rhodamine

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B), pesticides (pentachlorophenol, alachlor, and atrazine), and antibiotics (tetracycline,

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chloramphenicol, and lincomycin) (SI Figure S4 and Table S2). It was interesting to found that all

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these contaminants could be degraded within 60 min, suggesting the high efficiency of

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α-FeOOH-HA/H2O2 system. Subsequently, we systematically investigated the Fenton oxidation

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mechanism of α-FeOOH-HA/H2O2 system with alachlor as the model contaminant. As shown in

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Figure 2, alachlor could not be degraded by H2O2, and only 6% of alachlor was degraded in the

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α-FeOOH/H2O2 system within 60 min, much lower than that (> 90%) in the α-FeOOH-HA/H2O2

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system. The low degradation efficiency of α-FeOOH/H2O2 system was attributed to the poor surface

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Fe(III)/Fe(II) cycle of α-FeOOH. Although HA of high concentration (10 mmol/L) could react with

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H2O2 to produce •OH,16 only about 5% of alachlor was degraded within 60 min during the

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HA/H2O2 process because the concentration of HA was as low as 0.5 mmol/L in this

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α-FeOOH-HA/H2O2 system. The influence of HA concentration on the alachlor degradation in the

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α-FeOOH-HA/H2O2 system was therefore investigated at the initial pH values of about 5.0. Along

261

with increasing the HA concentration from 0.0 mmol/L to 0.75 mmol/L, the alachlor degradation rate

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first increased, but then decreased slightly when the HA concentration was changed from 0.75 11



mmol/L to 1.0 mmol/L (SI Figure S5). The slight decrease of alachlor degradation rate at higher HA

264

concentration might be attributed to the competitive consumption of reactive oxygen species by

265

excess HA. As the degradation percentage of alachlor (92%) in case of 0.5 mmol/L HA was close to

266

that (94%) of 0.75 mmol/L HA, 0.5 mmol/L was chosen as the HA concentration for the subsequent

267

investigation.

268

The generation of •OH in the α-FeOOH-HA/H2O2 system was first investigated by adding •OH

269

scavenger. As shown in Figure 3a, the addition of iso-propanol (IPA) significantly decreased the

270

alachlor degradation efficiency of α-FeOOH-HA/H2O2 system, suggesting that •OH was the major

271

reactive species in the α-FeOOH-HA/H2O2 system. Meanwhile, the alachlor degradation curve did

272

not change when we bubbled Ar gas to remove dissolved oxygen from the α-FeOOH-HA/H2O2

273

system (Figure 3b), ruling out the participation of molecular oxygen in the alachlor degradation.

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Subsequently, ESR spectra were used to investigate the generation of •OH in different systems with

275

employing 5,5-dimethyl-l-pyrroline-N-oxide (DMPO) as the spin trapper (Figure 3c). The

276

DMPO-•OH adducts were only detected in the α-FeOOH-HA/H2O2 system, but not in the HA/H2O2

277

and α-FeOOH/H2O2 systems, revealing that HA could promote the •OH generation in the

278

α-FeOOH-HA/H2O2 system. Then, we compared the •OH generation rate (V•OH) in the

279

α-FeOOH-HA/H2O2 and α-FeOOH/H2O2 systems with using the reaction between BA and •OH (SI

280

Figure S6).5, 28, 29 The specific surface area of α-FeOOH was 33.2 m2/g (SI Figure S7). It was found

281

that the specific surface area normalized •OH generation rate constant ((1.1 ± 0.1) × 10-4 g s-1 m-2) of

282

the α-FeOOH-HA/H2O2 system was 104 times that ((3.7 ± 0.2) × 10-7 g s-1 m-2) of the

283

α-FeOOH/H2O2 system (Figure 3d) and 103-106 times those of other Fenton systems based on

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iron-containing materials (Table S3),5 further confirming the indispensable roles of HA in the

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enhanced •OH generation of the α-FeOOH-HA/H2O2 system.

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The roles of HA in the •OH generation of the α-FeOOH-HA/H2O2 system were thus investigated 12


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systematically. First, the pH variation associated with the addition of HA was checked. It was found

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that the addition of HA into the reaction solution of alachlor, α-FeOOH and H2O2 changed the pH

289

value of from 6.6 to 5.0. During the alachlor degradation in the α-FeOOH-HA/H2O2 system, the pH

290

value slightly changed from 5.0 to 5.3 within 60 min (SI Figure S8a), suggesting the H+

291

consumption and/or the OH- generation. Then, we investigated the alachlor degradation in the

292

α-FeOOH-HA/H2O2 system at different initial pH values, and found that alachlor could be degraded

293

faster in the α-FeOOH-HA/H2O2 system under pH 3.0 than pH 5.0, while the alachlor degradation

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percentage decreased from 90% to 55% when changing the initial pH value from 5.0 to 8.0 (SI

295

Figure S8b). In spite of this pH dependent phenomenon, the alachlor degradation percentage (55%)

296

in the α-FeOOH-HA/H2O2 system at pH 8.0 was significantly higher than that (~ 0%) in the absence

297

of HA, suggesting that HA could still promote the alachlor degradation under neutral and weak

298

alkaline conditions. Although •OH was found to be the main active species for the alachlor

299

degradation in the α-FeOOH-HA/H2O2 system under pH 5.0, the alachlor degradation mechanism in

300

the α-FeOOH-HA/H2O2 system under higher pH conditions was not clear. Hug et al. demonstrated

301

that As(III) might be oxidized by alternative oxidants, such as ferryl or other Fe(IV) species under

302

neutral pH, rather than •OH, in Fenton process.30 To verify the main species for alachlor degradation

303

in the α-FeOOH-HA/H2O2 system under higher pH (pH 6, 7 and 8), more •OH quenching

304

experiments with IPA were conducted (SI Figure S9). The addition of IPA completely inhibited the

305

alachlor degradation in the α-FeOOH-HA/H2O2 systems at different pH, confirming that •OH mainly

306

contributed to the alachlor removal in the α-FeOOH-HA/H2O2 system.

307

As aforementioned, surface Fe(III) of α-FeOOH could be reduced by HA. Regarding that

308

dissolved iron ions were not detected in the α-FeOOH-HA system or in the α-FeOOH-HA/H2O2

309

system in the presence/absence of alachlor at pH 5.0 with ICP-OES (SI Figures S1 and S10), we

310

concluded the H2O2 decomposition should mainly occur on the α-FeOOH surface. To understand the 13



Fe(III)/Fe(II) cycle on the α-FeOOH surface in the α-FeOOH-HA/H2O2 system, we employed

312

high-resolution X-ray photoelectron spectroscopy (HRXPS) to check the generation of Fe(II) on the

313

α-FeOOH surface during this Fenton process (Figure 4a). Previous study reported that the binding

314

energies of Fe2p3/2 and Fe2p1/2 XPS peaks of α-FeOOH were located at 711.0 and 724.9 eV,

315

respectively.31 In this study, the Fe2p3/2 and Fe2p1/2 peaks of the α-FeOOH samples treated with H2O

316

or H2O2 were also located at 711.0 and 724.9 eV, respectively. This consistence suggested that H2O

317

and H2O2 did not affect the Fe speciation on α-FeOOH surface. However, the Fe2p1/2 peak of the

318

α-FeOOH sample separated from the α-FeOOH-HA/H2O2 system appeared at 724.3 eV, with a

319

binding energy shift of -0.6 eV, suggesting that Fe(II) was generated on the α-FeOOH surface.32 To

320

further identify the generation of Fe(II) on the α-FeOOH surface separated from the

321

α-FeOOH-HA/H2O2 system, XPS depth profiling technique was employed (Figure 4a). As expected,

322

the binding energies of Fe2p1/2 peaks were located at 724.3, 724.4, 724.5 and 724.9 eV for the depths

323

of 0, 0.5, 1.0, 1.5 nm, respectively. Therefore, Fe(II) existed at least between 0 and 1.0 nm depth of

324

the α-FeOOH sample separated from the α-FeOOH-HA/H2O2 system. These results strongly

325

confirmed that Fe(II) was generated inside or on the surface of α-FeOOH in the α-FeOOH-HA/H2O2

326

system.

327

Subsequently, we checked the contribution of surface Fe(II) bound on α-FeOOH to the alachlor

328

degradation by chelating the surface Fe(II) with 1, 10-phenanthroline or bipyridine during the

329

alachlor degradation.33,

330

presence of phenanthroline or bipyridine and HA (SI Figure S11), and did not detect any dissolved

331

iron ions in the α-FeOOH/phenanthroline or bipyridine, α-FeOOH-HA/phenanthroline or bipyridine,

332

and α-FeOOH-HA/H2O2/phenanthroline or bipyridine systems at pH 5.0 within 1 h. These results

333

were reasonable because the reductive dissolution of goethite would cost a longer time under these

334

conditions. Interestingly, the surface Fe(II) chelation with 1, 10-phenanthroline or bipyridine

34

Meanwhile, we checked the iron dissolution from α-FeOOH in the

14


Page 14 of 28

Page 15 of 28


335

completely

inhibited

the

alachlor

degradation

and

the

H2O2 decomposition

in

the

336

α-FeOOH-HA/H2O2 system, confirming the indispensable roles of surface Fe(II) in the alachlor

337

removal with the α-FeOOH-HA/H2O2 system (SI Figure S12). Since Fe(II) on the α-FeOOH surface

338

was the main species to catalyze the H2O2 decomposition, H2O2 might be decomposed on the

339

α-FeOOH surface and the surface-bound •OH was generated in the α-FeOOH-HA/H2O2 system.

340

To confirm the decomposition of H2O2 on the α-FeOOH surface, the changes of surface hydroxyl

341

groups (-OH) and surface peroxide groups (-O22-) of α-FeOOH in the α-FeOOH/H2O2 and

342

α-FeOOH-HA/H2O2 systems were monitored with O1s XPS spectra analysis (Figure 4b). The O1s

343

peaks at 530.0 and 531.0 eV were attributed to the lattice oxygen (Fe-O) and lattice hydroxyl groups

344

(Fe-OH lattice) of α-FeOOH, respectively.31 Differently, the O1s XPS spectra of α-FeOOH separated

345

from the α-FeOOH/H2O2 system possessed another shoulder at 532.1 eV, which was ascribed to the

346

-O22- groups.35 As expected, this shoulder from -O22- groups on the α-FeOOH surface decreased and

347

the peak of surface hydroxyl groups at 533.3 eV increased obviously for α-FeOOH in the

348

α-FeOOH-HA/H2O2

349

surface-bound •OH via the H2O2 decomposition catalyzed by Fe(II) in situ generated on the

350

α-FeOOH surface in the presence of HA.

system.31 These phenomena

strongly confirmed the generation of

351

To further confirm the existence of surface bound •OH generated by the surface bound Fe(II) on

352

α-FeOOH, we compared the amounts of • OH generated in the supernatants of the

353

α-FeOOH-HA/H2O2 and α-FeOOH-HA/H2O2/F- systems with DMPO trapped ESR technique

354

(Figure 5), regarding that F- in the solution could desorb •OH bound on α-FeOOH surface by

355

forming strong hydrogen bond (surface-bound •OH···F-FeOOH).36, 37 As expected, the addition of F-

356

improved the intensity of DMPO trapped •OH ESR signal by 30% in the α-FeOOH-HA/H2O2

357

system after 120 s of reaction, well validating that the surface-bound •OH was generated by the

358

surface bound Fe(II) on α-FeOOH in the α-FeOOH-HA/H2O2 system. 15



Subsequently, we compared the α-FeOOH-HA/H2O2 system with the corresponding homogeneous

360

Fenton ones. Interestingly, the alachlor degradation percentage in the α-FeOOH-HA/H2O2 (92%) and

361

HA/Fe(II)/H2O2 (27%) systems were far higher than that (8%) in the Fe(II)/H2O2 system when the

362

soluble Fe(II) concentrations in these homogeneous Fenton ones were equal to the surface Fe(II)

363

concentration in the α-FeOOH-HA/H2O2 system (SI Figure S13), confirming that HA could promote

364

Fe(III)/Fe(II) cycle either in the solution or on the surface, and the α-FeOOH surface bound Fe(II)

365

was more active than the dissolved Fe(II) on the H2O2 decomposition, consistent with the previous

366

reports.38, 39 As expected, the α-FeOOH-HA/H2O2 system could generate more •OH than these two

367

homogeneous Fenton ones (SI Figure S14). Although the alachlor degradation rate in the

368

α-FeOOH-HA/H2O2 system was lower than that in the Fe(II)/H2O2 one after increasing the initial

369

concentration of Fe(II) to 1.1 mmol/L, the same as the concentration of α-FeOOH (SI Figure S15),

370

the α-FeOOH-HA/H2O2 system still succeeded with the reusability of α-FeOOH (SI Figure S16).

371

Surprisingly, the reactivity of α-FeOOH even slightly increased along with repeated use, further

372

confirming the indispensible roles of surface Fe(II) in the surface Fenton alachlor degradation.

373

The variations of nitrate ions, chloride ions and small molecule acids concentrations were then

374

monitored to evaluate the alachlor degradation in the α-FeOOH-HA/H2O2 system. Given that both

375

alachlor and HA contributed to the NO3- generation during the alachlor degradation in the

376

α-FeOOH-HA/H2O2 system, we therefore compared the differences of NO3- concentration changes

377

in the absence or presence of alachlor, which were 0.42 and 0.47 mmol/L, respectively. Their

378

difference (0.05 mmol/L) suggested about 68% of nitrogen mineralization of alachlor (SI Figure

379

S17a). During the alachlor degradation in the α-FeOOH-HA/H2O2 system, the Cl- concentration

380

increased to 0.072 mmol/L within 60 min, corresponding to 98% of dechlorination. Meanwhile, the

381

concentrations of HCOOH, CH3COOH and HOOCCOOH respectively increased to 37.3, 96.2 and

382

108.8 µmol/L after 60 min, revealing that about 43% of alachlor was degraded to small molecule 16


Page 16 of 28

Page 17 of 28


383

acids (SI Figure S17b). Moreover, the mineralization of alachlor in the α-FeOOH-HA/H2O2 system

384

was analyzed by total organic carbon (TOC) analysis (SI Figure S18a). After 5 h of reaction, about

385

35% of TOC was removed. When the H2O2 concentration was increased from 1 to 10 mmol/L, the

386

TOC removal percentage improved to 84% within 5 h. To eliminate the interference effect of H2O2,

387

the TOC content changes of alachlor solution in the presence and absence of H2O2 were also

388

monitored (SI Figure S18b). We found that the TOC content of alachlor solution was nearly

389

unchanged in the presence of H2O2, suggesting the TOC analysis of alachlor solution was not

390

affected by H2O2.

391

The degradation intermediates of alachlor (1) were then detected with LC-MS and GC-MS. Eight

392

degradation

intermediates,

393

2-chloro-N-(2,6-diethylphenyl)acetamide

394

N-(2-ethyl-6-methylphenyl)acetamide

395

methylbenzenamine (7), 1,3-diethyl-2-nitrosobenzene (8), 2-ethyl-6-methylphenol (9), were

396

identified

397

2,6-diethyl-N-methylbenzenamine (10) and N-(2-acetyl-6- ethylphenyl)-2-chloroacetamide (11),

398

were checked out by LC-MS (SI Figure S20). The mass spectra of all these detected intermediates of

399

alachlor were consistent with those reported in the literatures,40-42 which were summarized in Table

400

S4 (Supporting Information). The appearance of these degradation intermediates suggested a

401

possible alachlor degradation pathway involving dechlorination, alkylic oxidation, hydroxylation

402

and dealkylation process in the α-FeOOH-HA/H2O2 system, as shown in Scheme 1. First, C-Cl bond

403

and C-N bond were likely broken up to generate the N-dealkylation intermediates (compounds 2, 3,

404

4) and the dechlorination intermediate (compound 5). The compound 3 could be further oxidized by

405

•OH to generate the compound 11. Meanwhile, para-position carbon or N-adjacent atom was

406

attacked by •OH to generate the compounds 6, 7 and 10. Subsequently, the amino side chain on the

by

GC-MS

including

(SI

2-chloro-N-(2,6-diethylphenyl)-N-methylacetamide (3),

2-chloro-N-(2-ethylphenyl)acetamide

(5),

Figure

2,6-diethylbenzenamine

S19),

while

17


two

(6),

other

(2), (4),

2-ethyl-6-

intermediates,


benzene ring would be oxidized to generate the compounds 8 and 9. Eventually, •OH could attack

408

the benzene ring to generate small-molecule organic acids and finally mineralize alachlor.

409

Meanwhile, it was interesting to find that HA could be degraded along with alachlor in the

410

α-FeOOH-HA/H2O2 system (SI Figure S21a), accompanying with 84% of HA mineralization into

411

NO3- (SI Figure S21b). Therefore, we conclude that this surface Fenton system is very

412

environmental friendly.

413

According to the above results, we would like to account for the high pollutant removal

414

performance of α-FeOOH-HA/H2O2 surface Fenton system as follows (Scheme 2). First, iron-HA

415

complexes (≡Fe(III)-HA) formed on the α-FeOOH surface. Then, the electron transfer from HA to

416

surface ferric iron (≡Fe(III)) would produce surface bound ferrous species (≡Fe(II)) to decompose

417

H2O2 for the generation of abundant •OH to degrade alachlor and HA, accompanying with the

418

formation of surface ≡Fe(III)-HA complex. Subsequently, the formed ≡Fe(III)-HA complex would

419

be reduced by HA to trigger another cycle of surface Fenton alachlor degradation.

420

Environmental Implications. In geochemical processes, the contaminants transformation dynamics

421

and the geochemical cycling of other redox−active elements are strongly affected by iron redox

422

cycling. Therefore, the understanding of interaction between goethite and reductants is of important

423

environmental implications. In this study, the effective degradation of alachlor in the surface Fenton

424

system (α-FeOOH-HA/H2O2) was ascribed to the efficient Fe(III)/Fe(II) cycle over the α-FeOOH

425

surface promoted by NH2OH. Moreover, the reactivity of α-FeOOH even slightly increased along

426

with repeated use, further confirming the indispensable roles of surface Fe(II) in the surface Fenton

427

degradation of alachlor. As goethite, reductants, and hydrogen peroxide widely exit in nature, similar

428

surface Fenton reaction will occur in the environment, affecting the transformation and conversion

429

of organic contaminants. Thus, the mechanistic investigation of surface Fenton system in this study

430

can provide new insights into the intrinsic relationship between iron cycling and the geochemical 18


Page 18 of 28

Page 19 of 28


431

cycling of other redox−active elements, and also offer a new strategy with iron cycling to realize

432

pollutant control and environmental remediation.

433 434

ACKNOWLEDGEMENTS

435

This work was supported by Natural Science Funds for Distinguished Young Scholars (Grant

436

21425728), National Science Foundation of China (Grant 51472100), the 111 Project (Grant

437

B17019), Self−Determined Research Funds of CCNU from the Colleges’ Basic Research and

438

Operation of MOE (Grant CCNU14Z01001), excellent doctorial dissertation cultivation grant from

439

Central China Normal University (Grant 2016YBZZ037), and the CAS Interdisciplinary Innovation

440

Team of the Chinese Academy of Sciences.

441 442

Supporting

Information

Available

443

Additional descriptions, figures, and tables as mentioned in the text. These materials are

444

available free of charge via the internet at http://pubs.acs.org.

445 446

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447

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degradation of 2,4,6-tribromophenol in a heterogeneous Fenton-like system with an iron-loaded

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549 550 23


Environmental Science & Technology 551 552

Figure Captions

553 554

Figure 1. (a) Changes of dissolved Fe(II) and Fe(III) concentration released from α-FeOOH in the

555

presence of HA under different atmosphere; the initial HA concentration was 5 × 10-4 mol/L; (b)

556

Changes of Fe(II) content of α-FeOOH as a function of HA concentration; the initial HA

557

concentrations were from 0 to 2 × 10-3 mol/L. The dosage of α-FeOOH was 0.1 g/L. The initial pH

558

was 5.0.

559 560

Figure 2. (a) Time profiles of the alachlor degradation in different systems. (b) Plots of -ln(C/C0)

561

versus time for the degradation of alachlor. The initial concentrations of alachlor, HA and H2O2 were

562

7.4 × 10-5, 5 × 10-4 and 1 × 10-3 mol/L, respectively; the dosage of α-FeOOH was 0.1 g/L; the initial

563

pH was 5.0.

24


Page 24 of 28

Page 25 of 28


564 565

Figure 3. (a) The degradation of alachlor in the absence and presence of iso-propanol (IPA) as the

566

•OH scavenger. (b) Time profiles of the alachlor degradation in the α-FeOOH-HA/H2O2 system in

567

the presence of air or Ar. (c) DMPO trapped ESR spectra in different systems. (d) The rate of

568

formation of •OH measured by the reaction with benzoic acid. The initial concentrations of alachlor,

569

IPA, benzoic acid, HA and H2O2 were 0.01, 7.4 × 10-5, 2 × 10-3, 5× 10-4 and 1 × 10-3 mol/L; The

570

dosage of α-FeOOH was 0.1 g/L; the initial pH was 5.0.

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Figure 4. Fe 2p XPS depth profiling spectra (a) and O1s XPS spectra (b) of the α-FeOOH samples

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in different reaction systems. 25



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Figure 5. (a) DMPO trapped ESR spectra in the α-FeOOH-HA/H2O2 and α-FeOOH-HA/H2O2/F-

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systems. (b) Time profiles of the ESR signal of DMPO-•OH generation. The initial concentrations of

577

HA, H2O2 and F- were 5 × 10-4, 1 × 10-3 and 2 × 10-3 mol/L, respectively; the dosage of α-FeOOH

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was 0.1 g/L; the initial pH was 5.0.

579 580

Scheme 1. Possible alachlor degradation pathway in the α-FeOOH-HA/H2O2 system.

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Scheme 2. Schematic illustration for the possible hydroxyl radical generation mechanism in the

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α-FeOOH-HA/H2O2 system.

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TOC Art Figure

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Hydroxylamine Promoted Goethite Surface Fenton ... - ACS Publications

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