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Protocatechuic Acid Promoted Alachlor Degradation in Fe(III)/H2O2 Fenton System Yaxin Qin, Fahui Song, Zhihui Ai, Pingping Zhang, and Lizhi Zhang Environ. Sci. Technol., Just Accepted Manuscript • Publication Date (Web): 11 Jun 2015 Downloaded from http://pubs.acs.org on June 11, 2015

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Environmental Science & Technology

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Protocatechuic Acid Promoted Alachlor Degradation in Fe(III)/H2O2 Fenton System

3

Yaxin Qin, Fahui Song, Zhihui Ai, Pingping Zhang, and Lizhi Zhang*

4

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

5

Environmental Chemistry, College of Chemistry, Central China Normal University, Wuhan 430079,

6

People’s Republic of China

1

7 8

RECEIVED DATE (to be automatically inserted after your manuscript is accepted if required

9

according to the journal that you are submitting your paper to)

10 11 12 13 14 15 16 17 18 19 20 21 22

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

ACS Paragon Plus Environment


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ABSTRACT In this study, we demonstrate that protocatechuic acid (PCA) can significantly

24

promote the alachlor degradation in the Fe(III)/H2O2 Fenton oxidation system. It was found that the

25

addition of protocatechuic acid could increase the alachlor degradation rate by 10000 times in this

26

Fenton oxidation system at pH = 3.6. This dramatic enhancement of alachlor degradation was

27

attributed to the complexing and reduction abilities of protocatechuic ligand, which could form

28

stable complexes with ferric ions to prevent their precipitation and also accelerate the Fe(III)/Fe(II)

29

cycle to enhance the •OH generation. Meanwhile, the Fe(III)/PCA/H2O2 system could also work

30

well at near natural pH even in case of PCA concentration as low as 0.1 mmol/L. More importantly,

31

both alachlor and PCA could be effectively mineralized in this Fenton system, suggesting the

32

environmental

33

chromatography-mass spectrometry to identify the degradation intermediates of alachlor and then

34

proposed a possible alachlor degradation mechanism in this novel Fenton oxidation system. This

35

study provides an efficient way to remove chloroacetanilide herbicides, and also shed new insight

36

into the possible roles of widely existed phenolic acids in the conversion and the mineralization of

37

organic contaminants in natural aquatic environment.

benignity

of

PCA/Fe(III)/H2O2

Fenton

system.

We

employed

gas

38 39

KEYWORDS: Protocatechuic acid; Fe(III)/Fe(II) cycle; Alachlor; Degradation; Fenton reaction

40 41

Introduction

42

Alachlor, as a widely used chloroacetanilide herbicide to control annual grasses and broadleaf weeds

43

in corn, sorghum, as well as soybeans, has been widely detected in soil, surface water and

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groundwater around the world.1-4 It is classified as a group B2 carcinogen by the Environmental

45

Protection Agency of Unite States with a maximum contaminant level of 2 ppb in drinking water.5,6 2


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The half-lives of alachlor in soil and water are over 70 and 30 days, respectively.7,8 Alachlor is also

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known as a highly toxic endocrine disrupting chemical. It may cause problems with eyes, liver,

48

kidneys, spleen, or experience anemia, even increase the risk of cancer.9-12 Therefore, it is of great

49

importance to develop effective and environmental friendly methods to remove alachlor.

50

As alachlor of high toxicity and chemical stability cannot be effectively degraded by conventional

51

biological remediation processes, scientists seek for advanced oxidation processes (AOPs) for the

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alachlor removal. These AOPs include ozonation, electrocatalysis oxidation, ultrasonic treatment,

53

and photocatalytic oxidation, etc.4-6,13-17 Although most of these methods can degrade alachlor to

54

some degree, the total organic carbon (TOC) removal efficiency of alachlor was not satisfied,

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indicative of its incomplete destruction. For example, the TOC removal of alachlor was merely less

56

than 10% after the treatment with high O3 dosages (14.1 mg/L) or O3/H2O2, even though it could be

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completely degraded.13 Therefore, it is still a great challenge to develop low cost and green methods

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to effectively mineralize alachlor.

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Among the AOPs, Fenton (Fe(II)/H2O2 or Fe(III)/H2O2) systems have been extensively studied in

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view of their high efficiency, simplicity and environmental friendliness.18,19 The classic Fenton

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system could generate hydroxyl radical via the decomposition of H2O2 catalyzed by Fe(II) (Eq. 1).

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Hydroxyl radical is highly reactive and can oxidize most organic substances rapidly and

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non-selectively.20-22 Unfortunately, ferric ions are quickly accumulated during the Fenton reactions,

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because the reaction rate of ferric ions and H2O2 is much slower than that of ferrous ions and H2O2

65

(Eq. 1 and 2). The formed ferric ions are inclined to precipitate as iron hydroxides (Eqs. 3-5), which

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would block the Fe(III)/Fe(II) cycle and thus slow the Fenton reactions.23,24 Obviously, it is of great

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importance to realize more effective Fe(III)/Fe(II) cycle during Fenton reactions for the oxidation of

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organic contaminants. 3



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Fe(II) + H 2 O 2 → Fe(III) + ⋅OH + OH −

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Fe(III) + H 2 O 2 → Fe(II) + ⋅HO 2 + H +

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Fe(II) + 2OH − → Fe(OH ) 2

(3)

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4Fe(OH) 2 + O 2 + 2H 2O → 4Fe(OH)3

(4)

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Fe(III) + 3OH − → Fe(OH ) 3

(5)

k = 63-76 M-1•s-1 k = (0.1-1.0) ×10-2 M-1•s-1

(1) (2)

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The UV light irradiation could accelerate the Fe(III)/Fe(II) cycle and thus improve the oxidation

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performance of Fenton system, but inevitably increases the cost and the energy consuming.25,26 It is

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known that aminopolycarboxylic acids such as ethylenediaminetetraacetic acid (EDTA), are able to

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prevent iron precipitation for more efficient Fenton oxidation of organic pollutants even at neutral

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pH. However, the wide application of EDTA may cause some adverse environmental consequences

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because of its poor biodegradability and strong heavy metal chelating ability.27,28 Recently, Wu et al.

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demonstrated the use of a new and strong complexing agent ethylenediamine-N,N'-disuccinic acid

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(EDDS) in the homogeneous Fenton process. They found that the performance of bisphenol A

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oxidation in the EDDS-driven Fenton reaction was much higher at near neutral or basic pH than at

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acidic pH owing to the formation of •HO2 or •O2- radicals and the presence of different forms of the

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complex Fe(III)-EDDS as a function of pH. EDDS could maintain iron in soluble forms, and also act

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as a superoxide radical-promoting agent to enhance the generation of Fe(II) and the production of

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•OH.29 However, the mineralization of bisphenol A and EDDS in the EDDS-driven Fenton reaction

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was not investigated. As a structural isomer of EDTA, EDDS might also suffer from high

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environmental risk for practical application.30-33

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Protocatechuic acid (PCA) is a typical phenolic acid widely existed in the edible plants, vegetables

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and fruits. It can be isolated from leaves of Ilex chinensis, which is an ingredient of some beverages

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and a Chinese herbal medicine.34,

35

It was reported that PCA was efficacious to against 4


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carcinogenesis or cardiovascular diseases,36, 37 and could be biodegraded.38, 39 Therefore, PCA is

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more environmentally benign than aminopolycarboxylic acids. Regarding that PCA possesses

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remarkable chelation ability with ferric ions and reduction ability to reduce Fe(III), 40-42 it could act

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as the complexing and reducing agent simultaneously to maintain iron in soluble forms and also

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accelerate the Fe(III)/Fe(II) cycle for the efficient Fenton oxidation. Moreover, the reaction rate

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constant (5.0 × 109 M-1•s-1) of •OH and PCA is relatively high, suggesting that PCA might be

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degraded along with organic contaminants, which can further reduce the environmental risk of

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residual organic chelators. In this study, we systemically investigate the effects of PCA on the

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Fe(III)/Fe(II) cycle and the alachlor oxidation performance of Fe(III)/H2O2 Fenton system. The fates

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of both alachlor and PCA during the PCA/Fe(III)/H2O2 Fenton oxidation were carefully studied with

102

gas chromatography-mass spectrometry and high-performance liquid chromatography with tandem

103

mass spectrometry. The purpose of this study aims to develop high efficient Fenton oxidation

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systems with environmentally benign characteristic.

105 106

Experimental Section

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Chemicals and Materials. Protocatechuic acid (PCA), Fe(NO3)3•9H2O, FeSO4•7H2O, alachlor,

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hydrogen peroxide (H2O2, 30% in water), acetic acid, hydroxylamine hydrochloride, sodium acetate

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(NaAc), tert-butyl alcohol (TBA), 1, 10-phenanthroline, H2SO4 and NaOH were all of analytical

110

grade and purchased from Sinopharm Chemical Reagent Co., Ltd. China. Alachlor was purchased

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from Sigma-Aldrich. Acetonitrile, acetone, and dichloromethane were of HPLC grade and obtained

112

from Merk KGaA. All chemicals were used as received without further purification. Deionized water

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was used in all experiments. H2O2 stock solutions were prepared by diluting 30% H2O2. Stock

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solutions of Fe(III) and Fe(II) with concentrations of 2 mmol/L were prepared by dissolving 5


Environmental Science & Technology 115

Fe(NO3)3•9H2O and FeSO4•7H2O in deionized water directly. The concentrations of protocatechuic

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acid and alachlor stock solutions were 2 mmol/L and 0.22 mmol/L, respectively. All the stock

117

solutions were prepared freshly for use. NaOH and H2SO4 solutions were used to adjust the pH value

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of the solutions.

119

Degradation Experiments. All the experiments were performed in 25 mL conical flask under

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constant magnetic stirring at normal temperature (25 ± 5 °C) and pressure. Typically, 10 mL of 0.22

121

mmol/L alachlor stock solution, 1 mL of 2 mmol/L PCA, and 1 mL of 2 mmol/L Fe(III) were mixed

122

with water to make the final volume of 20 mL. H2O2 was then added to trigger the degradation

123

experiments. The initial pH value of the degradation solution was 3.6 without pH adjusting. 1 mL of

124

the degradation solutions were sampled out with using a pipettor at regular intervals and 200 µL of

125

ethanol was added immediately to the sampled solution to quench the reaction for the subsequent

126

high pressure liquid chromatography (HPLC) measurement. For comparison, the alachlor

127

degradation was also investigated in the Fe(II)/H2O2 or Fe(II)/PCA/H2O2 Fenton system. The

128

FeSO4•7H2O solution (pH = 5.4) was prepared freshly without pH adjusting. The initial pH value of

129

the Fe(II)/H2O2 Fenton system was adjusted to 3.6 with 1 mol/L H2SO4 solution. To check the role of

130

molecular oxygen on the alachlor degradation process, high-purity argon gas was pumped into the

131

solution for 30 min prior to the initiation of the reaction and during the entire degradation

132

experiment at a rate of 1.5 L/min to remove molecular oxygen.

133

TOC Measurement. The degradation of 0.11 mmol/L alachlor was chosen to study the

134

mineralization of organic molecules in the Fe(III)/PCA/H2O2 system. The initial concentrations of

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Fe(III), PCA and H2O2 were 0.1, 0.1, and 8 mmol/L, respectively. The experiment was carried out in

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50-mL conical flask under constant magnetic stirring. At predetermined intervals, 4 mL of solution

137

was sampled and analyzed immediately after filtration through a 0.22 mm membrane. 320 µL of 1 6


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mol/L H2O2 was added to the reaction system at each 80 min for the continuous H2O2 supply. For

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comparison, the TOC changes in the Fe(III)/PCA/H2O2 system without adding alachlor and the

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Fe(II)/H2O2 Fenton system were measured. The total organic carbon (TOC) content was determined

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by a Shimadzu TOC-V CPH analyzer. The anions were detected by using an ion chromatograph (IC,

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Dionex ICS-900, Thermo) equipped with an AS23 column. A mixture containing 0.8 mmol/L

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NaHCO3 and 4.5 mmol/L Na2CO3 was used as the mobile phase running at a flow rate of 1.0

144

mL/min. The injective volume was 10 µL.

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Analytical Methods. The concentration of alachlor was monitored by high performance liquid

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chromatography (HPLC, LC-20AT, Shimadzu) with an Agilent TC-C18 reverse phase column. The

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injection volume was 10 µL. The eluent contained 40% of 0.75% (w/w) acetic acid and 60% of

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acetonitrile. The flow rate was 1 mL/min. The UV detector was set at 225 nm and the temperature of

149

column was maintained at 40 °C.

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Ferrous ions concentration was measured with a modified 1,10-phenanthroline method. Typically,

151

0.5 mL of sample solution, 0.5 mL of 1 g/L 1, 10-phenanthroline, and 0.5 mL of 10% NaAc were

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mixed with water to make the final volume of 2 mL and then analyzed by a UV–vis

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spectrophotometer (UV-2550, Shimadzu, Japan). The absorption of Fe(II)-1,10-phenanthroline

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complex was measured at λ = 510 nm. Hydroxylamine hydrochloride was utilized as the reducing

155

agent for the measurement of the total iron ions concentration.

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Isothermal titration calorimetry (ITC) experiments were performed on an ITC200 machine

157

(MicroCal, USA) at 25 °C. Fe(II), Fe(III), and PCA with concentrations of 1, 1, and 10 mmol/L were

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prepared with initial pH of 3.6. The solutions were filtered through a 0.22 µm membrane and

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de-gassed before use. Typically, 2 µL of PCA solution was injected into the sample cell containing

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200 µL Fe(II) or Fe(III) with an interval of 100 s. 7



The possible degradation intermediates of alachlor were identified by gas chromatography-mass

162

spectrometry (GC-MS, Trace 1300 equipped with ISQ, Thermo Fisher Scientific, USA) with a DB-5

163

column (size 30 m × 0.25 mm) and high-performance liquid chromatography with tandem mass

164

spectrometry (LC-MS/MS, TSQ Quantum Access MAX, Thermo Fisher Scientific, USA). For

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pre-treatment, 20 mL of samples was extracted with 20 mL of dichloromethane for three times. The

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combined extracts were dried with anhydrous sodium sulfate and the dichloromethane was removed

167

using a rotary evaporators. The residue was dissolved in 1 mL of acetone for GC-MS detection or 1

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mL of methanol for LC-MS/MS detection. 1.0 µL of final extract was injected into GC-MS to

169

analyze the intermediate products. Solvent delay at 5 min and scan range from m/z 50 to 350 were

170

used. The oven temperature was first kept at 50 °C for 1 min, then elevated to 180 °C with a rate of

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30 °C min-1 and held for 10 min, from 180 to 280 °C with a rate of 10 °C min-1, and finally held at

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280 °C for 10 min. The flow rate of the carrier gas helium was 1.0 mL/min. The LC-MS/MS

173

analyses were conducted in the positive ionization mode on a Hypersil GOLD column. The gradient

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elution was regulated by water and acetonitrile at a flow rate of 0.5 mL/min as follows. 95% water

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and 5% acetonitrile isocratic elution were first used for 2 min, followed by a linear gradient to 70%

176

acetonitrile within 30 min, and hold at 70% acetonitrile for 10 min. The MS capillary temperature

177

was 270 °C and the spray voltage was 3.2 kV. The injection volume was 1.0 µL.

178 179

Results and Discussion

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UV-vis spectrophotometer was first used to investigate the chelation of PCA with ferrous and ferric

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ions. PCA had two characteristic absorption peaks at 255 and 289 nm. With the addition of ferric

182

ions, a new broad band in the range 450–650 nm and a significant red shift of the absorption peaks at

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225 and 289 nm were observed (Figure S1 in the SI), which were attributed to the chelation of Fe(III) 8


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with PCA. When ferrous ions were added to the PCA solution, the characteristic absorption peaks

185

did not shift significantly and the characteristic absorption peaks at 255 decreased slightly,

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suggesting that PCA might not chelate with Fe(II). Isothermal titration calorimetry (ITC) was then

187

utilized to measure the complexation constant of Fe(III)-PCA, which was found to be (2.7 ± 0.3) ×

188

105 M-1 (Figure S2a and Table S1 in the SI). The ITC characterization also confirmed that PCA

189

could not form complexes with Fe(II) (Figure S2b in the SI).

190

Figure 1a shows the degradation curves of alachlor versus time under different conditions. We

191

found that neither Fe(III)/PCA nor PCA/H2O2 could effectively degrade alachlor. Although Fe(III)

192

could react with H2O2 to produce Fe(II) and hydroperoxide radicals, the alachlor concentration did

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not obviously change in the absence of PCA within 5 min, and then only decreased 9.8% even with

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prolonging the reaction time to 240 min (Figure S3 in the SI). This was because the catalytic

195

decomposition of H2O2 by Fe(III) was very slow. Interestingly, 97.6% of alachlor was degraded in

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the Fe(III)/PCA/H2O2 system at the initial pH of 3.6 within 5 min, suggesting that the

197

Fe(III)/PCA/H2O2 system was highly efficient to remove alachlor. Both of the alachlor degradation

198

curves in the Fe(III)/H2O2 and Fe(III)/PCA/H2O2 Fenton system were found to fit

199

pseudo-second-order kinetics equations (Figures S3 and S4 in the SI). The apparent alachlor

200

degradation rate constant (55.9 ± 2.4 L•mmol-1•min-1) in the Fe(III)/PCA/H2O2 system was about

201

10000 times that ((40.6 ± 2.9) × 10-4 L•mmol-1•min-1) in the Fe(III)/H2O2 system. As

202

ethylenediaminetetraacetic acid (EDTA) and ethylenediamine-N,N'-disuccinic acid (EDDS) are

203

commonly studied aminopolycarboxylic compounds for Fenton reaction, we systematically

204

compared

205

EDDS/Fe(III)/H2O2 systems with the initial pH value of 3.6. It was found that the alachlor

206

concentration did not obviously decrease within 5 min in the EDTA/Fe(III)/H2O2 and

the

alachlor

degradation

in

the

PCA/Fe(III)/H2O2,

9


EDTA/Fe(III)/H2O2,

and


EDDS/Fe(III)/H2O2 systems (Figure S5a in the SI), and only 9.6% and 11.8% of alachlor were

208

degraded in the EDTA/Fe(III)/H2O2 and EDDS/Fe(III)/H2O2 systems even with prolonging the

209

reaction time to 120 min (Figure S5b in the SI). The comparison revealed that EDTA and EDDS

210

could not accelerate the Fe(III)/Fe(II) cycle to promote the Fenton oxidation of alachlor although

211

they could form complexes with iron ions. As the generation of Fe(II) is the rate-limiting step in the

212

Haber-Wiess cycle, the dramatic alachlor degradation enhancement in this study might be attributed

213

to the efficient Fe(III)/Fe(II) cycle in the presence of PCA.

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In order to check the roles of PCA on the Fe(III)/Fe(II) cycle, we studied the alachlor degradation

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in the Fe(II)/H2O2 Fenton system in the absence or presence of PCA. The initial pH value of

216

Fe(II)/H2O2 Fenton system was adjusted to 3.6 with H2SO4. It was found that 34.8% of alachlor was

217

decomposed in the Fe(II)/H2O2 Fenton system within the first 0.5 min and then the concentration of

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alachlor degradation did not change during the subsequent 4.5 min (Figure 2a), suggesting the

219

termination of Fe(II)/H2O2 Fenton reactions because of the rapid exhaustion of Fe(II) and the

220

unsatisfactory Fe(III)/Fe(II) cycle after 0.5 min. Interestingly, the presence of PCA could also greatly

221

enhance the alachlor degradation in the Fe(II)/H2O2 Fenton system. Moreover, the alachlor

222

degradation curve in the Fe(II)/PCA/H2O2 system was almost the same as that of the

223

Fe(III)/PCA/H2O2 system (Figure 2b). These results revealed the presence of PCA could accelerate

224

the Fe(III)/Fe(II) cycle and thus more efficiently catalyze the decomposition of H2O2 to produce

225

more •OH for the alachlor oxidation. In this study, H2O2 was decomposed by Fe(II) to produce

226

hydroxyl radical in both Fe(II)/PCA/H2O2 and Fe(III)/PCA/H2O2 systems. In the Fe(II)/PCA/H2O2

227

system, the concentration of Fe(II) decreased in the first 1 min because the consumption rate of Fe(II)

228

was higher than its generation rate (Figure 3a). As for the Fe(III)/PCA/H2O2 system, the

229

concentration of Fe(II) increased quickly in the first 1 min, indicating that the generation rate of 10


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Fe(II) was higher than its consumption rate. However, the consumption rate of Fe(II) was equal to its

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generation rate in both of the two systems after 1 min because the presence of PCA could promote

232

the efficient Fe(III)/Fe(II) cycle. Therefore, the generation of Fe(II) in both of Fe(II)/PCA/H2O2 and

233

Fe(III)/PCA/H2O2 systems may not be the rate-limiting step, resulting in their similar alachlor

234

degradation curves, although the consumptions of Fe(II) and Fe(III) were different in these two

235

systems. Therefore, we detected the Fe(II) concentration changes in the Fe(III)/PCA system (Figure

236

S6a in the SI) and found that the generation rate constant of Fe(II) in the Fe(III)/PCA system was

237

279 ± 6 M-1•s-1 (Figure S6b and Table S1 in the SI), much higher than that (63 - 76 M-1•s-1) of Fe(II)

238

and H2O2. These results suggested that the presence of PCA guaranteed the sufficient supply of Fe(II)

239

in both the Fe(II)/PCA/H2O2 and Fe(III)/PCA/H2O2 systems, accounting for their similar alachlor

240

degradation curves.

241

To further check the roles of PCA on the Fe(III)/Fe(II) cycle, we monitored the concentration

242

changes of dissolved ferrous and total iron ions during the alachlor degradation in the different

243

systems (Figure 3). For the Fe(II)/H2O2 Fenton system, the concentrations of ferrous and total iron

244

ions decreased quickly and the ferrous ions concentrations approached to a minimum value in the

245

first 0.5 min, confirming the exhaustion of ferrous ions in the absence of PCA. The ferrous and total

246

iron ions concentration change curves matched well with the alachlor degradation curve in the

247

Fe(II)/H2O2 Fenton system, as shown in Figure 2a. In the Fe(III)/H2O2 system, the concentration of

248

total iron ions decreased quickly, while the ferrous ions concentration was very low. This

249

phenomenon was attributed to the weak ferric ions reduction ability of H2O2 (Eq. 2). As soon as PCA

250

was added, the total iron ions concentration in both Fe(II)/PCA/H2O2 and Fe(III)/PCA/H2O2 systems

251

did not significantly change because of the PCA’s chelation with ferric ions. More interestingly, the

252

addition of PCA allowed the instantaneous appearance of ferrous ions in the Fe(III)/PCA/H2O2 11



system and then the concentration of ferrous ions kept relatively constant. Obviously, the

254

regeneration of Fe(II) was no longer the rate-limiting step of Fenton reactions in the presence of

255

PCA. These results strongly confirmed the indispensable chelating and reduction roles of PCA on

256

the effective Fe(III)/Fe(II) cycle to guarantee the steady supply of ferrous iron, which favored the

257

H2O2 decomposition to produce more hydroxyl radical for the alachlor degradation.

258

Since the presence of PCA could prevent the precipitation of iron ions and efficiently accelerate

259

the Fe(III)/Fe(II) cycle, the molar ratio of PCA to Fe(III) may strongly affect the alachlor

260

degradation in the Fe(III)/PCA/H2O2 system. As expected, the alachlor degradation rate constant

261

increased from 11.7 ± 2.7 to 55.9 ± 2.4 L•mmol-1•min-1 along with increasing the PCA concentration

262

from 0.05 to 0.1 mmol/L (Figure S7a and Table S2 in the SI). However, the alachlor degradation

263

began to decrease when the PCA concentration further increased to 0.2 mmol/L because PCA would

264

compete with alachlor for the reactive oxygen species consumption, as the reaction rate constant (7.0

265

× 109 M-1•s-1) of •OH with alachlor is only about 1.4 times that (5.0 × 109 M-1•s-1) with PCA.43, 44

266

Although PCA would consume the reactive oxygen species, the simultaneous degradation of PCA is

267

of great significance to lower the environmental risk caused by the residual organic chelators, which

268

could form complexes with various metal ions and cause adverse environmental consequences.

269

As the efficient Fe(III)/Fe(II) cycle in the Fe(III)/PCA/H2O2 system guarantee the steady supply of

270

ferrous ions, the production of reactive oxygen species and the subsequent alachlor degradation

271

strongly depended on the H2O2 concentration, which would be consumed via the Eq. 1 or 2

272

continuously. We subsequently investigated the effects of H2O2 dosage on the degradation of

273

alachlor (Figure S7b in the SI). The alachlor degradation rate increased significantly from 18.8 ± 1.7

274

to 55.9 ± 2.4 L•mmol-1•min-1 when the H2O2 concentration increased from 2 to 8 mmol/L (Table S2

275

in the SI). The slight decrease of alachlor degradation rate at higher H2O2 dosage may be attributed 12


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to the consumption of hydroxyl radical by the excess hydrogen peroxide.

277

Besides the catalytic decomposition of hydrogen peroxide, ferrous ions can also be oxidized by

278

molecular oxygen, which is thought to be an undesirable competitive Fe(II) consumption. We

279

therefore compared the aerobic and anaerobic degradation of alachlor to check the influence of

280

molecular oxygen on the redox cycle of Fe(III)/Fe(II) and the alachlor degradation in the

281

Fe(III)/PCA/H2O2 system. It was found that the removal of molecular oxygen with bubbling

282

high-purity Ar did not significantly affect the alachlor degradation (Figure S8a in the SI). The

283

oxidation rate constant of Fe(II) by molecular oxygen in the presence of PCA was found to be 0.033

284

h-1 (Figure S8b in the SI). As shown in Figure S8b, only about 11.6% of Fe(II) was oxidized to

285

Fe(III) in the presence of PCA at pH = 3.6 within 240 min, suggesting that the oxidation of

286

Fe(II)-PCA by molecular oxygen could not significantly influence the alachlor degradation in this

287

study. It is well known that ferrous ions could decompose H2O2 to produce •OH through Fenton

288

reaction, we employed iso-propanol as the •OH scavenger to probe the contribution of •OH to the

289

alachlor degradation and found that the addition of isopropanol induced the complete suppression of

290

the alachlor degradation (Figure S9 in the SI). This phenomenon revealed that •OH was the major

291

oxidative species for the alachlor degradation in the Fe(III)/PCA/H2O2 system. We thus conclude

292

that PCA does not change the type of reactive oxygen species, although it can promote the redox

293

cycle of Fe(III)/Fe(II) in the Fe(III)/PCA/H2O2 system.

294

On the basis of the above results and analyses, we proposed a possible mechanism to explain the

295

PCA promoted alachlor degradation in the Fe(III)/H2O2 Fenton system. First, PCA reacted with

296

Fe(III) to form a Fe(III)-PCA complex. Meanwhile, some ferric ions were reduced to ferrous ions by

297

PCA because of its relatively strong reduction ability. Subsequently, PCA could effectively enhance

298

the Fe(III)/Fe(II) cycle and prevent the generation of iron sludge to maintain enough ferrous ions in 13



the Fe(III)/PCA/H2O2 system, and thus favor the fast decomposition of H2O2 to produce highly

300

reactive •OH for the final oxidation of alachlor.

301

The influence of initial pH values on the alachlor degradation in the Fe(III)/PCA/H2O2 system was

302

then investigated (Figure S10 in the SI). The results revealed that PCA enhanced the alachlor

303

degradation at acid pH (pH < 7) because PCA could function as the complexing and reducing

304

agent to maintain iron in soluble forms and also quickly reduce Fe(III) to Fe(II) at acidic pH.

305

However, the degradation efficiency of alachlor significantly decreased along with increasing the

306

initial pH value from 3 to 7. We therefore compared the alachlor degradation in PCA/Fe(III)/H2O2,

307

EDTA/Fe(III)/H2O2, and EDDS/Fe(III)/H2O2 systems at near natural pH (Figures S11 and S12 in the

308

SI). The results indicated that Fe(III)/PCA/H2O2 could also work well at a near natural pH of 6.2 and

309

it was even more efficient than Fe(III)/EDDS/H2O2 at pH = 6.2 when the complexes concentration

310

was low (e. g. 0.1 mmol/L). When pH ≥ 7, PCA could not enhance the alachlor degradation because

311

the rate of the reaction between PCA and Fe(III) was highly dependent on pH, as reflected by the

312

decreased equilibrium concentrations of Fe(II) in the Fe(III)/PCA system along with increasing pH

313

(Figure S13a in the SI). Therefore, the superiority of PCA gradually disappeared along with

314

increasing pH since the Fe(III) complexing ability of PCA (logK = 5.4) was not as strong as those of

315

EDTA (logK = 25.1) and EDDS (logK = 22),45 so the weak complexing ability of PCA would result

316

in the quick precipitation of Fe(III) to nullify the reductive ability of PCA at neutral and alkaline

317

conditions (Figure S13b in the SI).

318

Metal ions are widely existed in nature and thus affect the Fenton oxidation of organic pollutants.

319

We therefore chose four kind of metal ions (Cu(II), Al(III), Ni(II), and Co(II)), which could form

320

complexes with PCA,46,

321

degradation in the Fe(III)/PCA/H2O2 (Figure S14 in the SI). When the concentrations of coexistent

47

to investigate the influence of coexistent metal ions on the alachlor

14


Page 14 of 28

Page 15 of 28


322

metal ions were equal to that of Fe(III), the addition of Ni(II) slightly increased the alachlor

323

degradation, but the presence of Cu(II), Al(III), Co(II) inhibited the alachlor degradation. Although

324

these three kinds of coexistent metal ions would inhibit the alachlor degradation, more than 80%

325

degradation percentage of alachlor could be still achieved in the Fe(III)/PCA/H2O2 system,

326

indicating that the Fe(III)/PCA/H2O2 system would be effective if the concentrations of coexistent

327

metal ions were equal to or smaller than that of Fe(III). Moreover, the complexes of these metal ions

328

and PCA could also be degraded owing to the environmental benignity of PCA.

329

Total organic carbon analysis was then employed to probe the fates of alachlor and PCA in the

330

Fe(III)/PCA/H2O2 system. As both alachlor and PCA contributed to the TOC values of the

331

Fe(III)/PCA/H2O2 system, we therefore compared the TOC changes in the Fe(III)/PCA/H2O2 system

332

in the presence or absence of alachlor. Obviously, the TOC changes in the Fe(III)/PCA/H2O2 system

333

in the absence of alachlor reflected the PCA mineralization, and the differences of TOC in the

334

Fe(III)/PCA/H2O2 system in the presence or absence of alachlor could be regarded as the TOC

335

removal of alachlor (Figure 4). The initial TOC values of the Fe(III)/PCA/H2O2 system in the

336

presence or absence of alachlor were 27.5 and 8.4 mg/L, respectively. Therefore, the initial TOC

337

value of alachlor was calculated to be 19.1 mg/L, consistent with the theoretical value (18.7 mg/L).

338

As shown in Figure 4, only 13.3% of alachlor and 32.2% of PCA in the Fe(III)/PCA/H2O2 system

339

were mineralized in the first 60 min, but the TOC values continued to decline with the second

340

addition of H2O2. About 42.5% of alachlor and 46.6% of PCA could be mineralized within 180 min

341

with continuously adding H2O2, suggesting the simultaneous mineralization of alachlor and PCA in

342

the Fe(III)/PCA/H2O2 system. In contrast, negligible TOC removal was observed during the alachlor

343

degradation in the Fe(II)/H2O2 Fenton system even in the presence of excess H2O2, although 34.8%

344

of alachlor could be degraded in this traditional Fenton system (Figure S15 in the SI). This 15



comparison revealed that protocatechuic acid could promote both the degradation and the

346

mineralization of alachlor in the Fenton system

347

Ion chromatograph were further used to monitor the concentration variations of released chloride

348

ions and small molecule acids in the Fe(III)/PCA/H2O2 system. The concentration of Cl- increased to

349

a maximum value of 0.135 mmol/L in the first 80 min, corresponding to 52.1% of dechlorination

350

(Figure 5). These results indicated that the alachlor degradation in the Fe(III)/PCA/H2O2 system

351

involved the cleavage of C-Cl bonds by •OH to release Cl-. Three kinds of small molecule acids

352

were identified, which were HCOOH, CH3COOH, and HOOCCOOH. The concentrations of

353

HCOOH, CH3COOH, and HOOCCOOH increased along with prolonging the degradation time

354

(Figure 5). The concentration changes of small molecule acids in the Fe(III)/PCA/H2O2 system in

355

the absence of alachlor were also measured. The concentrations of CH3COOH and HOOCCOOH

356

increased to 0.078 and 0.074 mmol/L after 220 min, suggesting that more than 43.4% of PCA was

357

converted to small molecule acids (Figure S16 in the SI). The TOC changes in the Fe(III)/PCA/H2O2

358

system in the absence of alachlor indicated that 46.6% of PCA was mineralized within 220 min

359

(Figure 4a).Therefore, we concluded that most (at least 90%) of PCA was converted to small

360

molecule acids, CO2, and H2O in the Fe(III)/PCA/H2O2 system within 220 min.

361

Meanwhile, the mass balance of carbon during the alachlor degradation in the Fe(III)/PCA/H2O2

362

system was also calculated. It was found that the Fe(III)/PCA/H2O2 system converted 43.7% of

363

alachlor and PCA to CO2, and more than 34.6% of alachlor and PCA to small molecule acids within

364

220 min, leaving over less than 21.7% of TOC. This value was significantly lower than the

365

theoretical TOC value (29.8%) of carbon atoms in the benzene ring of alachlor and PCA, suggesting

366

the successful benzene ring cleavage during the alachlor degradation in the Fe(III)/PCA/H2O2

367

system. 16


Page 16 of 28

Page 17 of 28


368

Gas chromatograph-mass spectrometry and high-performance liquid chromatography with tandem

369

mass spectrometry were used to investigate the alachlor degradation intermediates to propose the

370

possible alachlor degradation pathway in the Fe(III)/PCA/H2O2 system at pH 3.6. Seven kinds of

371

alachlor degradation intermediates were detected by gas chromatograph-mass spectrometry, which

372

included 2-chloro-N-(2,6-diethylphenyl) acetamide (m/z = 225), 2,6-diethyl-N-(methoxymethyl)

373

aniline (m/z = 193), 2,6-diethylaniline (m/z = 149), 3,5-diethyl-4- amino-phenol (m/z = 165),

374

1,3-diethylbenzene (m/z = 134), 2,6-diethylnitrosobenzene (m/z = 163), and 3-ethylbenzaldehyde

375

(m/z = 134) (Figure S17 in the SI). Meanwhile, three kinds of alachlor degradation intermediates

376

were identified by high-performance liquid chromatography with tandem mass spectrometry,

377

including 2-chloro-N-(2,6-diethyl-phenyl)-N-methyl-acetamide (m/z = 238), N-(2-acetyl-6-ethyl-

378

phenyl)-2-chloro-acetamide (m/z = 240), and (2,6-diethyl-phenyl)-methyl-amine (m/z = 162) (Figure

379

S18 in the SI). On the basis of these intermediates, we proposed a possible degradation pathway of

380

alachlor in the Fe(III)/PCA/H2O2 system (Scheme 1). First, the C-N bond was easily cleaved to form

381

N-dealkylation

382

2,6-diethyl-N-(methoxymethyl)aniline, and 2,6-diethylaniline. Subsequently, hydroxyl radical

383

attacked N-adjacent or para-position carbon atom to produce 3,5-diethyl-4-amino-phenol,

384

1,3-diethylbenzene, and 2-amino-3-hydroxybenzoic acid. The amine groups on the benzene ring

385

could be oxidized to produce 2,6-diethylnitrosobenzene and 4-nitrosophenol. Then, the substitution

386

of nitroso-groups on the benzene ring would produce 1,3-diethylbenzene. Finally, benzene ring was

387

cleaved under the attack of hydroxyl radical, leading to the formation of small-molecule organic

388

acids and inorganic species. Unfortunately, the intermediate products of PCA was not detected with

389

GC-MS probably because of their low concentration and/or instability.

390

Environmental

degradation

products

Implications.

such

Different

as

from

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

the

traditional

17


organic

acetamide,

chelators

like


aminopolycarboxylic acids, PCA is one kind of phenolic acids widely existed in many plants and

392

thus be of more environmental benignity. The addition of PCA could effectively promote the Fenton

393

degradation of alachlor because PCA can form complexes with Fe(III) to prevent the precipitation of

394

Fe(III), and also realize the effective Fe(III)/Fe(II) cycle. More importantly, both alachlor and PCA

395

can be effectively mineralized in the Fe(III)/PCA/H2O2 system, which can further reduce the

396

environmental risk caused by the residual aminopolycarboxylic acid chelators. Meanwhile, as H2O2

397

and Fe(III) are ubiquitous in natural water environment in spite of tiny amount, Fe(III)/H2O2 Fenton

398

reaction might take place naturally in various aquatic environment and thus participate in the

399

transformation of organic contaminants. Therefore, the widely existed phenolic acids are highly

400

possible to promote the conversion and the mineralization of organic contaminants in natural aquatic

401

environment via Fenton oxidation processes.

402 403

Acknowledgements. This work was supported by National Natural Science Funds for Distinguished

404

Young Scholars (Grant 21425728), National Science Foundation of China (Grants 21173093,

405

21177048, 21273088, and 21477044), Key Project of Natural Science Foundation of Hubei Province

406

(Grant 2013CFA114), and Self-Determined Research Funds of CCNU from the Colleges’ Basic

407

Research and Operation of MOE (Grant CCNU14Z01001 and CCNU14KFY002).

408 409

Supporting Information Available: Additional figures as mentioned in the text are available free of

410

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

411 412

References

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TiO2 suspensions. Environ. Sci. Technol. 2003, 37, 2310-2316.

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for carbon tetrachloride degradation in modified Fenton’s systems. Environ. Sci. Technol. 2004, 38,

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composite cathode for wastewater treatment. J. Hazard. Mater. 2009, 164, 18-25.

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(23) Ma, J. H.; Song, W. J.; Chen, C. C.; Ma, W. H.; Zhao, J. C.; Tang, Y. L. Fenton degradation of

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organic compounds promoted by dyes under visible irradiation. Environ. Sci. Technol. 2005, 39,

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5810–5815.

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(24) Kopf, S. H.; Henny, C.; Newman, D. K. Ligand-enhanced abiotic iron oxidation and the effects

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of chemical versus biological iron cycling in anoxic environments. Environ. Sci. Technol. 2013, 47,

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2602-2611.

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(25) Nie, Y.; Hu, C.; Qu, J.; Zhou, L.; Hu, X. Photoassisted degradation of azo dyes over

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FeOxH2x-3/Fe0 in the presence of H2O2 at neutral pH values. Environ. Sci. Technol. 2007, 41,

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4715-4719.

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(27) Nam, S.; Renganathan, V.; Tratnyek, P. G. Substituent effects on azo dye oxidation by the

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FeIII–EDTA–H2O2 system. Chemosphere 2001, 45, 59-65.

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(28) Englehardt, J. D.; Meeroff, D. E.; Echegoyen, L.; Deng, Y.; Raymo, F. M.; Shibata, T. Oxidation

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of aqueous EDTA and associated organics and coprecipitation of inorganics by ambient

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iron-mediated aeration. Environ. Sci. Technol. 2006, 41, 270-276.

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Fe(III)–EDDS complex in Fenton-like processes: from the radical formation to the degradation of

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Biodegradation of metal-[S, S]-EDDS complexes. Environ. Sci. Technol. 2001, 35, 1765-1770.

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(34) Benitez, F. J.; Beltran-Heredia, J.; Acero, J. L.; Gonzalez, T. Degradation of protocatechuic acid

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by two advanced oxidation processes: Ozone/UV radiation and H2O2/UV radiation. Water Res. 1996,

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Photo-Fenton treatment of water containing natural phenolic pollutants. Chemosphere 2003, 50, 22


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protocatechuic acid: a genetic analysis. Res. Microbiol. 2001, 152, 861-872.

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F. Structure-property-activity relationship of phenolic acids and derivatives. Protocatechuic acid

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alkyl esters. J. Agr. Food Chem. 2010, 58, 6986-6993.

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Verhe, R. Iron-chelation properties of phenolic acids bearing catechol and galloyl groups. Food

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Chem. 2006, 98, 23-31.

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protocatechuate 4,5-dioxygenase-nitrosyl complex. Evidence for direct substrate binding to the

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active site Fe2+ of extradiol dioxygenases. J. Biol. Chem. 1986, 261, 2170-2178.

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iron(III). Inorg. Chem. 1988, 27, 4563-4566.

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(45) Orama, M.; Hyvonen, H.; Saarinen, H.; Aksela, R. Complexation of [S,S] and mixed

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stereoisomers of N,N′-ethylenediaminedisuccinic acid (EDDS) with Fe(III), Cu(II), Zn(II) and Mn(II)

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ions in aqueous solution. J. Chem. Soc., Dalton Trans. 2002, 24, 4644-4648.

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O-labeled substrate and inhibitors to

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531

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535 536 537 538 539 540 541 542 543 544 545 546 547 548 549 550 551

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Page 25 of 28 552

Environmental Science & Technology Figure Captions

553 554

Figure 1. (a) Time profiles of the alachlor degradation in different systems; (b) Plots of 1/Ct versus

555

time for the alachlor degradation in the Fe(III)/PCA/H2O2 system. The initial concentrations of

556

alachlor, Fe(III), PCA and H2O2 were 0.11, 0.1, 0.1 and 8 mmol/L, respectively. The initial pH value

557

of the systems was 3.6.

558 559

Figure 2. Time profiles of the alachlor degradation in the system of (a) Fe(III)/PCA/H2O2 and

560

Fe(II)/H2O2; (b) Fe(III)/PCA/H2O2 and Fe(II)/ PCA/H2O2. The inset is the plots of 1/Ct versus time

561

for the alachlor degradation in the Fe(III)/PCA/H2O2 and Fe(II)/PCA/H2O2 systems. Initial

562

concentrations of alachlor, Fe(II), Fe(III), PCA and H2O2 were 0.11, 0.1, 0.1, 0.1 and 8 mmol/L,

563

respectively. The initial pH value of the systems was 3.6.

25



564 565

Figure 3. Variation of total iron ions concentration in different systems during the alachlor

566

degradation: (a) ferrous ions; (b) total iron ions. Initial concentrations of Fe(II), Fe(III), PCA and

567

H2O2 were 0.1, 0.1, 0.1 and 8 mmol/L, respectively. The initial pH value of the systems was 3.6.

568 569

Figure 4. (a) Time profiles of TOC removal in the Fe(III)/PCA/H2O2 system with and without

570

alachlor. (b) The TOC changes of alachlor in the Fe(III)/PCA/H2O2 system calculated by subtracting

571

the TOC values of the Fe(III)/PCA/H2O2 system in the absence of alachlor from those of the

572

Fe(III)/PCA/H2O2 system in the presence of alachlor. The initial concentrations of alachlor, Fe(III),

573

PCA and H2O2 were 0.11, 0.1, 0.1 and 8 mmol/L, respectively. The initial pH value of the systems

574

was 3.6.

26


Page 26 of 28

Page 27 of 28


575 576

Figure 5. Variation of chloride ions and small-molecule organic acids concentrations in the

577

Fe(III)/PCA/H2O2 system. The initial concentrations of alachlor, Fe(III), PCA and H2O2 were 0.11,

578

0.1, 0.1 and 8 mmol/L, respectively. The initial pH value of the systems was 3.6.

579 580

Scheme 1. A possible alachlor degradation pathway in the Fe(III)/PCA/H2O2 system.

581 582 583

27



TOC Art Figure

585

28


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H2O2

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