Comparison of Nitrilotriacetic Acid and [S, S]-Ethylenediamine-N, N

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Comparison of nitrilotriacetic acid and [S,S]-ethylenediamineN,N’-disuccinic acid in UV-Fenton for the treatment of oil sands process-affected water at natural pH Ying Zhang, Nikolaus Klamerth, Pamela Chelme-Ayala, and Mohamed Gamal El-Din Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b03050 • Publication Date (Web): 02 Sep 2016 Downloaded from http://pubs.acs.org on September 19, 2016

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Environmental Science & Technology

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Comparison of nitrilotriacetic acid and [S,S]-ethylenediamine-N,N’-

2

disuccinic acid in UV-Fenton for the treatment of oil sands process-

3

affected water at natural pH

4

Ying Zhang, Nikolaus Klamerth, Pamela Chelme-Ayala, Mohamed Gamal El-Din*

5 6 7

Department of Civil and Environmental Engineering, University of Alberta, Edmonton, AB, Canada T6G 1H9

8 9 10

Intended for: Environmental Science & Technology

11 12 13 14

*Corresponding Author:

15

Mohamed Gamal El-Din, Ph.D., P.Eng.

16

Professor, NSERC Senior Industrial Research Chair in Oil Sands Tailings Water Treatment

17

Department of Civil and Environmental Engineering

18

University of Alberta

19

Edmonton, Alberta, Canada T6G 1H9

20

Tel: +1 780 492 5124

21

Fax: +1 780 492 0249

22

Email: [email protected]

23 24 1 ACS Paragon Plus Environment


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ABSTRACT

26

The application of UV-Fenton processes with two chelating agents, nitrilotriacetic acid (NTA)

27

and [S,S]-ethylenediamine-N,N’-disuccinic acid ([S,S]-EDDS), for the treatment of oil sands

28

process-affected water (OSPW) at natural pH was investigated. The half-wave potentials of

29

Fe(III/II)NTA and Fe(III/II)EDDS and the UV photolysis of the complexes in MilliQ water and

30

OSPW were compared. Under optimum conditions, UV-NTA-Fenton exhibited higher efficiency

31

than UV-EDDS-Fenton in the removal of acid extractable organic fraction (66.8% for the former

32

and 50.0% for the latter) and aromatics (93.5% for the former and 74.2% for the latter).

33

Naphthenic acids (NAs) removals in the UV-NTA-Fenton process (98.4%, 86.0%, and 81.0% for

34

classical NAs, NAs + O (oxidized NAs with one additional oxygen atom), and NAs + 2O (oxidized

35

NAs with two additional oxygen atoms), respectively) under the experimental conditions were much

36

higher than those in the UV-H2O2 (88.9%, 48.7%, and 54.6%, correspondingly) and NTA-Fenton

37

(69.6%, 35.3%, and 44.2%, correspondingly) processes. Both UV-NTA-Fenton and UV-EDDS-

38

Fenton processes presented promoting effect on the acute toxicity of OSPW towards Vibrio

39

fischeri. No significant change of the NTA toxicity occurred during the photolysis of Fe(III)NTA;

40

however, the acute toxicity of EDDS increased as the photolysis of Fe(III)EDDS proceeded.

41

NTA is a much better agent than EDDS for the application of UV-Fenton process in the

42

treatment of OSPW.

43

44

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1. Introduction

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The bitumen extraction from oil sands and subsequent froth treatment process generate oil

49

sands process-affected water (OSPW), which is highly saline and acutely toxic to aquatic

50

organisms. Organic compounds in OSPW include unrecovered bitumen (oil and grease),

51

naphthenic acids (NAs), polyaromatic hydrocarbons (PAHs), BTEX (benzene, toluene, ethyl

52

benzene, and xylenes), and other organic compounds such as fulvic and humic acids.1-5 The

53

toxicity of OSPW is primarily attributed to the presence of NAs, which have biological half-lives

54

exceeding 13 years.6 NAs are a group of aliphatic and alicyclic carboxylic acids with a general

55

formula of CnH2n+zOx, where n is the carbon number, x is equal to two for classical NAs (c-NAs)

56

and three or more for oxidized NAs (oxy-NAs) formed after the oxidation of c-NAs,7 and z (zero

57

or a negative even integer) indicates the hydrogen deficiency due to the presence of rings or

58

double bonds.4 Advanced oxidation processes (AOPs) have been proposed as a complementary

59

approach to the biological treatment of OSPW due to their ability to degrade recalcitrant NAs

60

and reduce the overall toxicity of OSPW towards selected organisms.8-10 Although Fenton and

61

photo-Fenton processes are very common AOPs and have been used to treat wastewaters,11, 12

62

they have not been applied in the OSPW treatment, as a low pH is needed to prevent the

63

precipitation of Fe.

64

Chelating agents have been employed to form complexes with iron to enable Fenton

65

reaction at high pH.13-19 Aminopolycarboxylic acids (APCAs) are comprised of several

66

carboxylate groups bound to one or more nitrogen atoms, and they can form very stable

67

complexes with many di- or tri-valent metallic ions.20 Nitrilotriacetic acid (NTA) and [S,S]-

68

ethylenediamine-N,N’-disuccinic acid ([S,S]-EDDS) are the most common aminopolycarboxylic

69

acids used in the modified Fenton processes nowadays. NTA has been reported by Sun and 3 ACS Paragon Plus Environment


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Pignatello16 as one of the most active chelates among 50 chelating agents used for the

71

decomposition of H2O2 and the degradation of 2,4-dichlorophenoxyactic acid. EDDS as a

72

structural

73

environmentally safe compared to EDTA.21-23 Its performance is equal to EDTA in most cases

74

and even much better under specific conditions.24-26

isomer

of

ethylenediaminetetraacetic

acid

(EDTA)

is

biodegradable

and

75

UV irradiation has been widely applied to increase the efficiency of the modified Fenton

76

processes,27-31 due to its ability to reduce Fe(III)L to Fe(II)L (L denotes NTA or EDDS in this

77

case) at high pH. APCAs alone do not absorb solar light above 250 nm, and their photolysis is

78

negligible. The complexes APCAs form with iron have been reported to exist as monomeric or

79

dimeric species in aqueous solution, absorb light up to 400 nm, and present rapid photochemical

80

degradation under UV irradiation.20, 32, 33 To date, no research has compared NTA and EDDS in

81

their physicochemical characteristics or in the efficiency of the modified Fenton processes using

82

these two agents.

83

In this study, the UV-NTA/EDDS-Fenton processes were applied on the treatment of OSPW

84

for the first time. The reduction mechanism of Fe(III)NTA/EDDS to Fe(II)NTA/EDDS was

85

discussed by comparing the redox potentials of the complexes with the potentials of H2O2/O2•-

86

and O2/O2•-. We investigated the decomposition of Fe(III)NTA and Fe(III)EDDS under UV

87

irradiation in MilliQ water and in OSPW, and discussed the influence of H2O2 on the

88

decomposition based on the reaction of •OH with chelating agents and the formation of binuclear

89

µ-peroxo adducts. The efficiency of the UV-NTA/EDDS-Fenton processes was investigated on

90

the removal of acid extractable organic fraction (AEF), aromatics and the overall toxicity of

91

OSPW towards Vibrio fischeri. The toxicity of NTA, EDDS, Fe(III)NTA/EDDS and their

92

degradation products towards Vibrio fischeri was measured and discussed based on the

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generation of ammonia, nitrite, and nitrate in the UV-NTA/EDDS-Fenton processes. The

94

removals of c-NAs and oxy-NAs in the UV-NTA-Fenton, UV-H2O2 and NTA-Fenton processes

95

were also presented and compared.

96

2. Materials and Methods

97

2.1 Chemicals and stock solutions

98

NaOH, H2O2 (30%), FeSO4·7H2O, 98% H2SO4, 1,10-phenanthroline, acetic acid, methylene

99

chloride, sodium acetate, ammonium hydroxide, and sodium nitrate were purchased from Fisher

100

Scientific Co. Canada. Nitrilotriacetic acid (99%), ethylenediamine-N,N’-disuccinic acid (35%

101

in water), bovine catalase, sodium tetraborate decahydrate, hydrochloric acid, dimethyl sulfoxide

102

(DMSO), and optima methanol and acetonitrile were purchased from Sigma Aldrich.

103

Tris(hydroxymethyl)aminomethane was purchased from Alfa Aesar. Titanium (IV) oxysulfate

104

was purchased from Fluka Analytical. Specific kits for nitrate (TNT 53), nitrite (TNT 839), and

105

ammonia (TNT 69) measurement were from HACH. Filters used were 0.45 µm nylon membrane

106

filters (GE Healthcare Life Sciences).

107

Raw OSPW with a pH value of 8.3 was collected from an active oil sands tailings pond in

108

Fort McMurray, Alberta, Canada, and was stored at 4˚C prior to use. The typical composition of

109

OSPW is provided in Table S1 in the Supporting Information (SI). 0.048 M Fe(II) stock solution

110

was prepared by dissolving FeSO4·7H2O in MilliQ water (Millipore Corporation) at pH 3 prior

111

to Fenton reactions. 0.18 M NTA stock solution was prepared by dissolving NTA in MilliQ

112

water at pH 4.34 0.18 M EDDS stock solution with a final pH of 9.13 was prepared by diluting 35%

113

EDDS solution with MilliQ water. It has been stated in the literature that Fe(III)EDTA complex

114

prepared by the oxidation of Fe(II)EDTA was more effective in some radical reactions than the

115

complex prepared directly from a ferric salt.35 In our study, therefore, we used Fe(II) source to



116

prepare Fe(III)NTA/EDDS complexes. The dose of iron for the test solutions (0.089 mM) was

117

chosen based on natural Fe concentration of 0.8-3 mg/L in OSPW.5 The dose of NTA/EDDS

118

(0.72 mM) chosen was much higher than that of Fe because that at low ratio of NTA/EDDS:Fe,

119

the fast photodecomposition of Fe-NTA/EDDS leads to the generation of less reactive Fe-

120

complexes and the precipitation of Fe, resulting in the low efficiency of the UV-NTA/EDDS-

121

Fenton processes. The results of the degradation of a model NA compound, cyclohexanoic acid

122

(CHA), at low ratio of NTA/EDDS:Fe in the NTA/EDDS-Fenton processes can be found in our

123

previous papers.36, 37

124

2.2 UV irradiation test

125

UV irradiation experiments were conducted by placing a 100-mL beaker (5.4-cm diameter)

126

with a 80 mL sample solution on a magnetic stirrer under a collimated beam UV apparatus

127

(Model PSI-I-120, Calgon Carbon Corporation, Pittsburgh, PA, USA) equipped with a 1-kw

128

medium-pressure (MP) Hg-lamp (Calgon Carbon, Pittsburgh, PA, USA), shown in Figure S1 in

129

the SI. The emission of the lamp was from 200 nm to 530 nm (Figure S2). NTA (or EDDS) and

130

Fe(II) stock solution were added to form Fe-NTA (or Fe-EDDS) complex, afterwards H2O2 was

131

added, and then the shutter of the warmed-up UV lamp was opened immediately. Because of the

132

natural buffering effect of OSPW, the pH value of OSPW did not change significantly (from 8.3

133

to 8.2) by adding NTA/EDDS and Fe(II) stock solution. Samples were taken periodically and

134

added drops of bovine catalase (1g/L) to destroy the excess of H2O2. Control experiments were

135

conducted under UV irradiation only. Experiments were carried out in duplicates.

136

2.3 Analytical methods

137

H2O2 was measured using titanium (IV) oxysulfate (DIN 38402H15) at 410 nm15, 38 and Fe

138

was measured using 1.10-phenanthroline (ISO 6332) at 510 nm39 (GENESYS™ 10S UV-Vis


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spectrophotometer). pH was controlled with 0.1 M NaOH and 0.1 M H2SO4 using a pH meter

140

(Fisher Scientific, AR 50). NTA and EDDS in the form of Fe(III)-complexes were analyzed by

141

HPLC-UV (Agilent Technologies, 1260 Infinity) with a column from Phenomenex (C18, 5 µm,

142

150 mm × 4.6 mm). Cyclic voltammetry was conducted with a Metrohm Autolab

143

electrochemical workstation using a three-electrode system. The measurement of metal ions in

144

OSPW was carried out using a Perkin Elmer’s Elan 6000. Nitrate, nitrite, and ammonia analysis

145

was conducted using a DR 3900 spectrometer (HACH) with the specific kits.

146

Fourier transform infrared (FT-IR) spectra were obtained by using a PerkinElmer Spectrum

147

100 FT-IR Spectrometer (PerkinElmer Life and Analytical Sciences, Woodbridge, ON, CA)

148

using the protocol described elsewhere.40,

149

recorded in a Varian Cary Eclipse fluorescence spectrometer (Mississauga, ON, Canada) with a

150

scanning speed of 600 nm/min and a photomultiplier (PMT) voltage of 800 mV. 1H nuclear

151

magnetic resonance (NMR) analysis was carried out with a 3-channel Agilent/Varian Unity

152

Inova Spectrometer (Agilent Technologies, Santa Clara, CA, USA). The analysis of NAs with an

153

ultra-performance liquid chromatography time-of-flight mass spectrometry (UPLC TOF-MS)

154

can be found in Wang et al.42 The acute toxicity of the samples towards Vibrio fischeri was

155

processed using Microtox○R 81.9% screening test with a Microtox○R 500 Analyzer by following

156

the manufacturer’s instructions.

157

41

Synchronous fluorescence spectra (SFS) were

Details of the analyses of NTA, EDDS, metal ion, cyclic voltammetry, FT-IR, NMR, TOF-

158

MS, and toxicity are provided in the SI.

159

3. Results and Discussions

160

3.1 Half-wave potentials



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Cyclic voltammograms of Fe(III/II)NTA and Fe(III/II)EDDS at pH 7 and at pHs 8 and 9 are

162

included in Figure 1a and Figure S3, respectively, with the oxidation and reduction peaks marked

163

with red circles. Potential values of Fe(III/II)NTA at pHs 7, 8, and 9 at the oxidation peaks were

164

200, 200, and 150 mV, respectively, and at the reduction peaks were -792, -900, and -1050 mV,

165

respectively. Potential values of Fe(III/II)EDDS at pHs 7, 8, and 9 at the oxidation peaks were

166

650, 270, and 190 mV, respectively, and at the reduction peaks were -750, -820, and -840 mV,

167

respectively. The E1/2 values (Figure 1b) were calculated by taking the average of the potentials

168

at the oxidation and reduction peaks and referring the average potentials (by adding 236 mV to

169

them) to the standard hydrogen electrode (SHE). The E1/2 of Fe(III/II)NTA at pHs 7, 8 and 9

170

were obtained as -64, -114, and -214 mV (vs SHE), respectively, and the E1/2 of Fe(III/II)EDDS

171

at the same pH values were 186, -39, and -89 mV (vs SHE), respectively.

172

HO2•/O2•- (with a pKa of 4.8) is mainly in the form of O2•- at high pH.43, 44 Redox potentials

173

of O2/O2•- and H2O2/O2•- at pH 7 were reported as 70 and 360 mV, respectively.45 Unfortunately,

174

the redox potentials of these species at higher pHs (such as 8 and 9) are not available in the

175

literature. By comparing the redox potentials of Fe(III/II)NTA (-64 mV) and Fe(III/II)EDDS

176

(186 mV) at pH 7 with that of H2O2/O2•- (360 mV), it can be concluded that it is impossible for

177

H2O2 to reduce Fe(III)NTA to Fe(II)NTA or Fe(III)EDDS to Fe(II)EDDS in the way shown in

178

Eq. 1.

179

‫)ܫܫܫ(݁ܨ‬L + ‫ܪ‬ଶ ܱଶ → ‫)ܫܫ(݁ܨ‬L + ‫ܱܪ‬ଶ• /ܱଶ•ି + ‫ ܪ‬ା

180

This conclusion is in agreement with a previous report that H2O2 could not reduce

181

complexed iron,35, 46, 47 and the reduction of Fe(III)L in the modified Fenton system is O2•--

182

dependent not H2O2-dependent.35 The redox potential of O2/O2•- at pH 7 (70 mV) makes the


(1)

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183

radical capable of reducing Fe(III)EDDS to Fe(II)EDDS; however, incapable of reducing

184

Fe(III)NTA to Fe(II)NTA in the way shown in Eq. 2.

185

‫ ܮ)ܫܫܫ(݁ܨ‬+ ‫ܱܪ‬ଶ• /ܱଶ•ି → ‫ ܮ)ܫܫ(݁ܨ‬+ ܱଶ

186

This hypothesis coincides with previous observation that no Fe(II) was detected in the NTA-

187

Fenton process at pHs 7 and 8.36, 48 Nonetheless, the regeneration of Fe(II) was observed in the

188

EDDS-Fenton process at pH 8.37 Therefore, with the presence of abundant chelates, the

189

decomposition of H2O2 in the EDDS-Fenton system should be faster than that in the NTA-

190

Fenton system due to the fast reaction between Fe(II)EDDS and H2O2, indicating a higher

191

generation rate of •OH in the former system.

192

3.2 Photodegradation of NTA and EDDS

(2)

193

The metal ion concentration in the raw OSPW and the stability constants of the metal-

194

NTA/EDDS complexes are included in Table S2. The most concentrated metal ions were Na

195

(600 mg/L), Mg (25 mg/L), K (29 mg/L), and Ca (41 mg/L). Li, B, Si, Sr, and Ba were in a range

196

of 0.1-4 mg/L. Most transition metal ions existed in trace amounts (< 0.05 mg/L), such as Ni

197

(0.009 mg/L) and Cu (0.010 mg/L), and the total Fe concentration was below the detection limit.

198

The stability constants of metal-chelates increases with increasing charge and decreasing

199

radius of the metal ions, and with the number of coordinating groups in the ligands.49 Generally,

200

the stability constants of metal-NTA complexes are smaller than those of metal-EDDS

201

complexes, such as 15.933, 50, 51 for Fe(III)NTA and 20.652 for Fe(III)EDDS, 18.4524, 53, 54 for

202

Cu(II)NTA and 14.255 or 13.149 for Cu(II)EDDS, and 7.449 for Mn(II)NTA and 8.9524, 53, 56 for

203

Mn(II)EDDS (ignoring the temperature (20 or 25 ºC) difference for these values), indicating a

204

higher strength of the interaction of metal ions in OSPW with EDDS than that with NTA. EDDS

205

with six coordinating sites (two N donors and four O donors) forms both five- and six-membered



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206

chelate rings with metal ions with a ratio of 1:1,57 while NTA as a tetradentate ligand with four

207

coordinating sites (one N donor and three O donors) can form 1:1 and 2:1 complexes with metal

208

ions.58 The lower number of coordinating sites of NTA compared to EDDS attributes to the

209

smaller stability constants of metal-NTA complexes.

210

The degradation of 0.72 mM NTA and EDDS, respectively, under UV irradiation and in the

211

UV photolysis of Fe(III)L (Fe = 0.089 mM) in MilliQ water and in OSPW showed no significant

212

difference (Figure 2). No NTA degradation and negligible EDDS degradation (< 4.0%) were

213

achieved under UV irradiation, which is consistent with the literature.20, 32, 33 81.0% and 84.6%

214

degradation of Fe(III)NTA and Fe(III)EDDS, respectively, occurred in the UV photolysis of

215

Fe(III)L within 30 min due to the fast photolysis of metal-APCA complexes.20,

216

decomposition of NTA takes place at C-N bond and generates iminodiacetic acid (IDA), glycine,

217

oxalic acid, ammonia, and carbon dioxide.33 No oxidation products of EDDS can be found in the

218

literature; however, the biodegradation of EDDS takes place with a C-N cleavage and gives

219

weak chelating agents N-(2-aminoethyl) aspartic acid (AEAA), fumarate, and eventually CO2.59

32, 33

The

220

The degradation of Fe(III)NTA and Fe(III)EDDS under UV irradiation in MilliQ water was

221

significantly accelerated by the addition of 5.88 mM H2O2, with a slightly faster degradation of

222

the former complex (97.5% degradation in 7 min) than that of the latter complex (97.8%

223

degradation in 10 min). The acceleration of the degradation of Fe-complexes in the presence of

224

H2O2 was primarily attributed to two main reasons: i) the interaction of •OH with NTA

225

(4.77±0.24 × 108 M-1s-1 36 at pH 8) and EDDS (2.48 ± 0.43×109 M-1s-1

226

proved to be the main radical responsible for the degradation of contaminants in the

227

NTA/EDDS-Fenton processes,36, 37 which are the dominated processes in the UV-NTA/EDDS-

228

Fenton processes. The results in this study do not exclude the formation of ferryl ions; however,


37

at pH 8). •OH was

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229

they are not the primary oxidative species under the experimental conditions; and ii) the

230

formation of binuclear µ-peroxo adducts, H2O2-Fe(III)NTA/EDDS. The second hypothesis was

231

made based on previous reports of H2O2-adducts for the reaction of H2O2 with Fe(III),60

232

Fe(III)EDTA,43,

233

binuclear µ-peroxo adduct contains singlet oxygen (1∆g) and exhibits a unique reactivity through

234

its direct interaction with organic substrates. The activity of the peroxide ion depends on the

235

complexes used.64 The µ-peroxo adducts can be converted to high-valence Fe(IV)-oxo species,

236

which is highly reactive towards abundant substrates.65 Based on these theories, the degradation

237

of Fe(III)L in the presence of H2O2 might be to some extent caused by the interaction with the

238

activated peroxide ion or by the oxidation with the Fe(IV)-oxo species. The attribution of the

239

•

240

by their reaction rate constants with •OH. Therefore, the higher H2O2 interference on the NTA

241

degradation should be primarily attributed to the formation of the H2O2-adduct.

61, 62

and Fe(III)NTA.63 The highly activated peroxide ion trapped in the

OH attack on the NTA degradation was very limited compared with that on EDDS, as suggested

242

It is very interesting to notice that the degradation of the Fe(III)NTA complex (0.089 mM

243

Fe and 0.72 mM NTA) under UV irradiation with or without the presence of H2O2 in OSPW was

244

much slower than that in MilliQ water. In OSPW, the Fe(III)NTA degradation in 30 min was

245

95.9% with the presence of 5.88 mM H2O2 and 56.4% without H2O2, while the Fe(III)NTA

246

degradation in MilliQ water was 97.5% in 7 min with the presence of 5.88 mM H2O2 and 81.0%

247

in 30 min without H2O2. Due to the interference of the water matrix, the total EDDS

248

concentration in OSPW could not be determined. However, the H2O2 decomposition (Figure S4b)

249

in the UV-EDDS-Fenton process for the treatment of OSPW slowed down after 15 min (for

250

example, 4.41 mM H2O2 decomposed to 1.46 mM in the first 15 min and then decomposed to

251

1.05 mM in the following 15 min), as shown in Figure S4. Moreover, no signal of Fe(III)EDDS



252

could be detected in the HPLC-UV after 15 min, indicating complete decomposition of EDDS at

253

that time, slower than that in MilliQ water (97.8% degradation in 10 min). Some organics in

254

OSPW, such as fulvic and humic acids,1-5 could form less reactive co-complexes with Fe(III)L.

255

The heavy metal ions in OSPW (Table S2) competed with iron for the chelating agents. The

256

higher interference of the metal ions on EDDS than that on NTA might primarily attribute to the

257

faster decomposition rate of EDDS in OSPW. Furthermore, the high UV absorption of OSPW

258

decreased the UV energy for the photolysis of Fe(III)L. These three interferences attributed

259

primarily to the hindering effect of OSPW on the photolysis rate of Fe(III)NTA/EDDS.

260

3.3 FT-IR results

261

Acid extractable organic fraction (AEF) in the raw and UV-NTA/EDDS-Fenton treated

262

OSPW was measured with FT-IR (Figure 3), and the spectra are provided in Figure S5. The O-H

263

stretch in the phenol-like structures registers at 3600 cm-1,41, 66 while aromatic and alkene C-H

264

stretches both occur over 3000 cm-1 with the peaks at 3100-3000 cm-1 reflecting the aromatic C-

265

H stretch. C=C aromatic stretches appear often in pairs, with one at 1615-1580 cm-1 and the other

266

one at 1510-1450 cm-1.67 The peaks between 670 and 900 cm-1 can be attributed to =C-H out-of-

267

plane in the aromatic compounds.67 The absorbance at 1743 cm-1 and 1706 cm-1 are for the

268

monomeric and dimeric forms of carboxylic groups, respectively, and they were applied for the

269

AEF calculation because that these values are commonly used by the oil sands industry to

270

evaluate the organic acid compounds in OSPW.41, 68, 69

271

In the UV-NTA-Fenton process, the AEF (64.52 mg/L) removal in 30 min (Figure 3)

272

increased with increasing H2O2 dose at H2O2 < 5.88 mM, and decreased with increasing H2O2

273

dose at H2O2 > 5.88 mM, with a highest removal of 66.8% at 5.88 mM H2O2 (H2O2

274

decomposition shown in Figure S4). In the UV-EDDS-Fenton process, AEF removal increased


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275

with increasing H2O2 dose at H2O2 < 4.41 mM, and decreased with increasing H2O2 dose at

276

H2O2 > 4.41 mM, with a highest removal of 50.0% at 4.41 mM H2O2. The lower AEF removal at

277

higher H2O2 dose in these two processes was due to the scavenging effect of H2O2 on •OH with a

278

reaction rate constant reported as 2.70×107 M-1s-1.70-72 The AEF in OSPW includes c-NAs, oxy-

279

NAs, and other organics containing carboxylic groups,68, 73 with the total NAs comprising about

280

50.0% of AEF.74, 75 The results indicate that UV-NTA-Fenton is more efficient in the reduction

281

of AEF in OSPW compared with UV-EDDS-Fenton, primarily due to the lower scavenging

282

effect of NTA on •OH than that of EDDS, as discussed in Section 3.2.

283

3.4 NMR and SFS results

284

The 1H NMR spectra of the raw and treated OSPWs were acquired to compare the removal

285

of organics in different processes (Figure 4a). The peaks at 11.4-12.5 ppm (I), 6.5-8.0 ppm (II),

286

and 0.3-2.2 ppm (III) are assigned to the hydrogen atoms in carboxylic groups, aromatic rings,

287

and aliphatic chains, respectively.42,

288

decreased by 61.5% (from 5.2 to 2.0), for peak II by 74.2% (from 3.1 to 0.8), and for peak III by

289

46.6% (from 99.5 to 53.1), while in the UV-NTA-Fenton process the same peaks were reduced

290

by 90.4% (from 5.2 to 0.5), 93.5% (from 3.1 to 0.2) and 80.4% (from 99.5 to 19.5), respectively.

291

In order to make the NMR spectra between 0 and 4 ppm more readable, the original 1H NMR

292

spectra for raw OSPW and OSPW treated with the UV-NTA/EDDS-Fenton processes are

293

provided in Figure S6 in the SI. In a study of OSPW treatment with 200 mg/L potassium

294

ferrate(VI) conducted by Wang et al.42, 29.7%, 44.4%, and 23.8% area decreases were achieved

295

for peaks I, II and III, respectively, much lower than those obtained in this study.

76

In the UV-EDDS-Fenton process, the area for peak I

296

SFS analysis was applied to measure fluorophore compounds, which are most likely

297

aromatic compounds present in OSPW. Peak I at 273 nm (Figure 4b) is the signal of one-ring



Page 14 of 32

298

aromatics and aromatics containing two or more unfused aromatic rings. Peak II at 310 nm and

299

peak III at 330 nm are the signals of aromatics with two and three fused rings, respectively.42, 77,

300

78

301

Peak II and III were completely removed. The removal of aromatics at Peak I increased with

302

increasing H2O2 dose, with no peak observed at H2O2 dose over 4.41 mM. In the UV-EDDS-

303

Fenton process with H2O2 from 1.47 to 8.82 mM (Figure 4c), aromatics at Peak I were

304

significantly removed, and the aromatics at Peak II and III were completely removed. The high

305

emission intensity at wavelengths 400-500 nm was related to the degradation products of

306

Fe(III)EDDS, as indicated by the SFS spectra of the photolyzed complex in Figure S7. A peak of

307

photolyzed Fe(III)EDDS in MilliQ water (Figure S7) showed up at the same wavelength (440

308

nm) as the peak at 1.47 mM H2O2 in Figure 4c. Unfortunately, detailed information of the new

309

peaks is not available in the literature. It was reported that one-ring aromatics have higher

310

stability than two- and three-ring aromatics towards oxidation.42 Aromatics with two or more

311

fused rings have higher electron density, leading to higher activity towards oxidation.42,

312

Moreover, aromatic compounds with greater size and increasing alkyl substitution have higher

313

quantum yields, resulting in higher sensitivity towards photochemical oxidation.80-82 The

314

production of one-ring aromatics from the decomposition of two- and three-ring aromatics also

315

explains the apparent stability of one-ring aromatics.42

In the UV-NTA-Fenton process (Figure 4b), with a H2O2 dose over 1.47 mM, the aromatics at

79

316

The reactions between •OH and organic matters usually take place by adding the radical to

317

C=C double bonds or by abstracting H-atoms from C-H bonds.83 The second-order rate constants

318

of the reaction of •OH with alkenes and aromatics are in a range of 109-1010 M-1s-1.84, 85 The •OH

319

attack on the aromatic compounds is preferentially by adding themselves to the aromatic rings,85,

320

86

and the intermediate adducts formed are in an excited state initially and transformed rapidly to


Page 15 of 32


321

a stabilized form, which might decompose back to the reactants or react further with scavenger

322

molecules, such as O2.86 Maki et al.87 reported a significant decline in the aromatic fraction and

323

an increase in the resin and asphaltene fractions after the sunlight irradiation of crude oil.

324

Photooxidation products of aromatic hydrocarbons were reported as corresponding alcohols,

325

aldehydes, ketones, and acids.81, 88, 89

326

Both 1H NMR and SFS spectra showed very high removal of aromatics in the UV-NTA-

327

Fenton process, in coincidence with the depleted signals of the aromatic C-H stretch at 3100-

328

3000 cm-1 in the FT-IR spectra (Figure S5), indicating high capability of this process to degrade

329

aromatics.

330

3.5 Toxicity results

331

Microtox ○R 81.9% screening test was applied to evaluate the efficiency of the UV-

332

NTA/EDDS-Fenton processes on the removal of the acute toxicity of OSPW towards Vibrio

333

fischeri. The inhibition effect of the raw OSPW was 39.5% (shown in Figure 5a at 0 mM H2O2),

334

close to 40-47% reported previously.80, 90 The inhibition effect of OSPW treated with the UV-

335

NTA-Fenton process increased to 70.8% at 1.47 mM H2O2, then decreased with increasing H2O2

336

dose, and finally reached 61.8% at 8.82 mM H2O2. In the UV-EDDS-Fenton process, the

337

inhibition effect of the treated OSPW increased gradually with increasing H2O2 dose from 39.5%

338

at 0 mM H2O2 to 72.0% at 8.82 mM H2O2. The increased inhibition effect of the photooxidation

339

treated OSPW agreed well with Shu et al.80 Maki et al.87 also reported an increase in the toxicity

340

against Artemia as the photooxidation of biodegraded oil proceeded. The photooxidation process

341

significantly increased the water-soluble fraction (medium-molecular-weight aromatic alcohols

342

and ethers) of oil, which can lead to a higher general toxicity.88,

343

fluorophore organic compounds in OSPW (Figure 4b-c) led to the production of alcohols,


91

The photolysis of the


344

aldehydes, ketones, and acids, which are considered to be responsible for the increased acute

345

toxicity towards Vibrio fischeri.80, 81, 88, 89

346

In order to test the acute toxicity of NTA, EDDS, Fe(III)NTA/EDDS and their products

347

towards Vibrio fischeri, the UV-NTA/EDDS-Fenton reactions were carried out in MilliQ water

348

with 5.88 mM H2O2, 0.089 mM Fe, and 0.72 mM NTA/EDDS. The degradation of NTA and

349

EDDS under these conditions is illustrated in Figure 2 with complete depletion of the agents at

350

12 min. NTA presented 9.4% inhibition effect (Figure 5b), indicating low acute toxicity of the

351

agent towards Vibrio fischeri. The acute and sub-acute toxicity of NTA has been reported to be

352

negligible,92 and the agent shows no adverse effects to aquatic life.93 No significant change of the

353

inhibition effect occurred during the photolysis of Fe(III)NTA with an effect value of 17.6% at

354

12 min (100% decomposition of Fe(III)NTA), indicating low acute toxicity of the Fe(III)NTA

355

products. EDDS and Fe(III)EDDS solution presented 12.7% and 12.1% inhibition effect towards

356

Vibrio fischeri, respectively (Figure 5b). The toxicity of the Fe(III)EDDS solution increased

357

gradually as the photolysis of the complex proceeded and reached 61.9% at 12 min (100%

358

decomposition of Fe(III)EDDS). EDDS degradation product fumarate is basically non-toxic.94

359

No data of the toxicity of AEAA is available in the literature.

Page 16 of 32

360

Ammonia (pKb = 4.75, in the main form of ammonium at the experimental pH 8.3)

361

generated in the UV-NTA/EDDS-Fenton processes was measured, as well as nitrite and nitrate.

362

Ammonium generated increased with time and achieved 3.0 and 5.4 mg/L at the end of the

363

reaction for NTA and EDDS processes, respectively. No nitrite and low concentration of nitrate

364

(0.5-1.4 mg/L) were achieved in both processes. The acute toxicity of 0.1-50 mg/L ammonium,

365

nitrate, and nitrite in the form of ammonium hydroxide, sodium nitrate, and sodium nitrite,

366

respectively, towards Vibrio fisheri was investigated. No inhibition effect of ammonium and


Page 17 of 32


367

nitrate, and very low inhibition effect of nitrite (< 6%) were observed, indicating the irrelevance

368

of these species to the acute toxicity towards Vibrio fisheri. Therefore, it can be concluded that

369

the high acute toxicity of the photolyzed Fe(III)EDDS was attributed to the production of AEAA.

370

NTA is a much better chelating agent than EDDS due to the higher organic removal in the

371

UV-NTA-Fenton process and lower toxicity risk of the agent towards aquatic organisms. The

372

contamination of NTA would not be an environmental concern because NTA is biodegradable

373

and the metal-NTA complexes can undergo rapid photodegradation.95, 96 Actually, only very low

374

level of NTA has been found in natural waters despite the use of NTA in industrial and consumer

375

products.93

376

3.6 NA removal

377

The overall percent distribution of NAs based on the carbon and -z number (hydrogen

378

deficiency) in the raw OSPW and in the OSPW treated with the UV-NTA-Fenton, UV-H2O2, and

379

NTA-Fenton processes are illustrated in Figure 6a-d, respectively. The distributions of c-NAs in

380

the raw OSPW with respect of carbon and –z number are provided in Table S3. The most

381

abundant species in the c-NAs were NAs with two (-z = 4) and three rings (-z = 6) (47.8% of the

382

total c-NAs), and c-NAs with carbon number ranging from 13-18 (81.0% of the total c-NAs) as

383

previously reported.97 The concentration of the total c-NAs in the raw OSPW was 25.4 mg/L,

384

and it decreased to 0.4 mg/L (98.4% removal), 2.8 mg/L (88.9% removal), and 7.7 mg/L (69.6%

385

removal) for the OSPW treated with the UV-NTA-Fenton, UV-H2O2, NTA-Fenton processes,

386

respectively. Only 12.8% NA removal was achieved by UV irradiation (data not shown here).

387

The initial concentration of NAs + O (oxy-NAs with one additional oxygen atom) in the raw

388

OSPW was 15 mg/L, and it decreased to 2.1 mg/L (86.0% removal), 7.7 mg/L (48.7% removal),

389

and 9.7 mg/L (35.3% removal) after the treatment with the UV-NTA-Fenton, UV-H2O2, and



Page 18 of 32

390

NTA-Fenton processes, respectively (Figure S8a-d). NAs + 2O (oxy-NAs with two additional

391

oxygen atoms) had an initial concentration of 16.3 mg/L in the raw OSPW, and it decreased to

392

3.1 mg/L (81.0% removal), 7.4 mg/L (54.6% removal), and 9.1 mg/L (44.2% removal) after the

393

treatment with the UV-NTA-Fenton, UV-H2O2, and NTA-Fenton processes, respectively (Figure

394

S9a-d). The UV-NTA-Fenton process exhibited higher removal of c-NAs (98.4%), NAs + O

395

(86.0%), and NAs + 2O (81.0%) in contrast with the UV-H2O2 and NTA-Fenton processes,

396

which was primarily attributed to the rapid reduction of Fe(III)NTA to Fe(II)NTA by UV light

397

and the fast reaction of Fe(II)NTA and H2O2,98 leading to the production of more •OH. The high

398

removal of c-NAs and oxy-NAs in the UV-NTA-Fenton process suggests the high efficiency of

399

this process on the degradation of total NAs in OSPW. Wang et al.42 reported removal of c-NAs,

400

NAs + O, and NAs + 2O in OSPW of 64.0%, -18.4%, and -18.0%, respectively, after the

401

treatment with 200 mg/L potassium ferrate(VI). The negative removal of NAs + O, and NAs +

402

2O was due to the reproduction of these species from the oxidation of c-NAs and the

403

transformation of the oxy-NAs at high n and z numbers to the oxy-NAs at low n and z numbers.

404

This explanation can also be used to illustrate the relative lower removal of oxy-NAs than that of

405

c-NAs in this study.

406

High hydrogen deficiency of NAs might be a function of the number of alicyclic rings or

407

due to the presence of aromatic rings or double bonds.66 However, the deficiency of the aliphatic

408

C-C double bonds in OSPW was confirmed with the NMR spectra, which showed no signal at

409

4.5-5.5 ppm assigned to the double bonds.42, 76 The •OH attack on the alicyclic rings in NAs

410

through H abstraction produces organic radicals, which then react with O2 to form peroxyl

411

radicals, as indicated by the reaction of •OH with CHA.99 The reaction between •OH and the

412

aromatic rings in NAs, if any, can be referred to the •OH attack on the aromatic compounds in


Page 19 of 32


413

Section 3.4, which are initiated predominantly through the radical addition to the aromatic rings.

414

The oxidation of NAs by •OH generated more soluble molecules (i.e., oxy-NAs with less carbon

415

number or fewer cyclic rings), as indicated by the increased solubility of NAs after oxidation.

416

4. Environmental Significance

417

This is the first application of the UV-chelate-modified Fenton on the OSPW remediation.

418

This work compared, for the first time, the physicochemical characteristics of NTA and EDDS,

419

the photolysis of Fe-NTA/EDDS in MilliQ water and in OSPW, and the efficiency of the UV-

420

NTA/EDDS-Fenton processes for the removal of organics and the OSPW toxicity towards Vibrio

421

fischeri. Compared to UV-EDDS-Fenton, UV-NTA-Fenton exhibited higher efficiency in the

422

removal of AEF and aromatics. NA removal in the UV-NTA-Fenton process was much higher in

423

contrast with the UV-H2O2 and NTA-Fenton processes. Overall, NTA is a much better chelating

424

agent than EDDS for the application of the UV-Fenton process on the treatment of OSPW at

425

natural pH. However, the acute toxicity of the OSPW treated with the UV-NTA-Fenton process

426

towards Vibrio fischeri increased compared to the raw OSPW, which warrants more research to

427

confirm the causes. The findings obtained in this study have significant impact for the further

428

development of this process and OSPW remediation.

429

Acknowledgement

430

This research was supported by a research grant from a Natural Sciences and Engineering

431

Research Council of Canada (NSERC) Senior Industrial Research Chair (IRC) in Oil Sands

432

Tailings Water Treatment through the support by Syncrude Canada Ltd., Suncor Energy Inc.,

433

Shell Canada, Canadian Natural Resources Ltd., Total E&P Canada Ltd., EPCOR Water

434

Services, IOWC Technologies Inc., Alberta Innovates - Energy and Environment Solution, and

435

Alberta Environment and Parks. The financial supports provided by Trojan Technologies and an



Page 20 of 32

436

NSERC Collaborative Research and Development (CRD) grant, and the Helmholtz-Alberta

437

Initiative are also acknowledged. The authors would like to thank the group of Dr. Ted Sargent

438

from the Department of Electrical and Computer Engineering at the University of Toronto for the

439

help in the measurements of half-wave potentials. Thanks also go to Dr. Rongfu Huang for NAs

440

analysis and Mr. Yuan Chen for his contribution in data mining, and Mr. Chengjin Wang for his

441

help in the NMR analysis.

442

Supporting Information

443

Details of the analytical methods and supplementary tables and graphs associated with this

444

article are provided in the Supporting Information. This material is available free of charge via

445

the Internet at http://pubs.acs.org.

446

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447 448 449 450 451 452 453 454 455 456 457 458 459 460 461 462 463 464 465 466 467 468 469 470

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49. Smith, R. M.; Martell, A. E., Critical stability constants, enthalpies and entropies for the formation of metal complexes of aminopolycarboxylic acids and carboxylic acids. Sci. Total. Environ. 1987, 64, (1-2), 125-147. 50. Cuculić, V.; Pižeta, I.; Branica, M., Voltammetry of Dissolved Iron (III)–Nitrilotriacetate–Hydroxide System in Water Solution. Electroanalysis 2005, 17, (23), 2129-2136. 51. Anderegg, G., The stability of iron (III) complexes formed below pH= 3 with glycinate, iminodiacetate, β-hydroxyethyliminodiacetate, N, N-di-(Hydroxyethyl)-glycinate, nitrilotriacetate and triethanolamine. Inorg. Chim. Acta 1986, 121, (2), 229-231. 52. Orama, M.; Hyvönen, H.; Saarinen, H.; Aksela, R., Complexation of [S, S] and mixed stereoisomers of N, N′-ethylenediaminedisuccinic acid (EDDS) with Fe (III), Cu (II), Zn (II) and Mn (II) ions in aqueous solution. J. Chem. Soc., Dalton Trans. 2002, (24), 4644-4648. 53. Martell, A. E.; Smith, R. M., Critical stability constants. Plenum Press: New York, 1974-1989; Vol. 1-6. 54. Gorelova, R.; Babich, V.; Gorelov, I., Potentiometric study of copper complex formation with ethylenediaminedisuccinic and ethylenediaminetetraacetic acids. ZHURNAL NEORGANICHESKOI KHIMII 1971, 16, (7), 1873-&. 55. Bolton, H.; Girvin, D. C.; Plymale, A. E.; Harvey, S. D.; Workman, D. J., Degradation of metalnitrilotriacetate complexes by Chelatobacter heintzii. Environ. Sci. Technol. 1996, 30, (3), 931-938. 56. Majer, J.; Jokl, V.; Dvorakova, E.; Juracova, M., New complexones. XIII. Potentiometric and electrophoretic study of ethylenediamine-N, N′-disuccinic acid and its metal chelates. Chem Zvesti 1968, 22, 415-423. 57. Pavelčík, F.; Majer, J., The crystal and molecular structure of lithium [(S, S)-N, N'ethylenediaminedisuccinato] cobaltate (III) trihydrate. Acta Crystallographica Section B: Structural Crystallography and Crystal Chemistry 1978, 34, (12), 3582-3585. 58. Anderegg, G., Critical survey of stability constants of NTA complexes. Pure and Applied chemistry 1982, 54, (12), 2693-2758. 59. Schowanek, D.; Feijtel, T. C.; Perkins, C. M.; Hartman, F. A.; Federle, T. W.; Larson, R. J., Biodegradation of [S, S],[R, R] and mixed stereoisomers of ethylene diamine disuccinic acid (EDDS), a transition metal chelator. Chemosphere 1997, 34, (11), 2375-2391. 60. Gallard, H.; De Laat, J.; Legube, B., Spectrophotometric study of the formation of iron (III)hydroperoxy complexes in homogeneous aqueous solutions. Water Res. 1999, 33, (13), 2929-2936. 61. Walling, C.; Kurz, M.; Schugar, H. J., Iron (III)-ethylenediaminetetraacetic acid-peroxide system. Inorg. Chem. 1970, 9, (4), 931-937. 62. Francis, K. C.; Cummins, D.; Oakes, J., Kinetic and structural investigations of [Fe III (edta)]–[edta= ethylenediaminetetra-acetate (4–)] catalysed decomposition of hydrogen peroxide. J. Chem. Soc., Dalton Trans. 1985, (3), 493-501. 63. Nishida, Y.; Goto, A.; Akamatsu, T.; Ohba, S.; Fujita, T.; Tokii, T.; Okada, S., Iron Chelates in Biological Systems. Its Relevance to Induction of Pathogenesis of Tissue Damage and Carcinogenesis. Chem. Lett. 1994, (3), 641-644. 64. Nishida, Y.; Nasu, M.; Akamatsu, T., Reaction between binuclear iron(III) compounds and DMPO (5, 5-dimethyl-3, 4-dihydropyrrole N-oxide). J. Chem. Soc., Chem. Commun. 1992, (2), 93-94. 65. Pierre, J.; Fontecave, M., Iron and activated oxygen species in biology: the basic chemistry. Biometals 1999, 12, (3), 195-199. 66. Reinardy, H. C.; Scarlett, A. G.; Henry, T. B.; West, C. E.; Hewitt, L. M.; Frank, R. A.; Rowland, S. J., Aromatic naphthenic acids in oil sands process-affected water, resolved by GCxGC-MS, only weakly induce the gene for vitellogenin production in zebrafish (Danio rerio) larvae. Environ. Sci. Technol. 2013, 47, (12), 6614-6620. 67. Coates, J., Interpretation of Infrared Spectra, A Practical Approach. In Encyclopedia of Analytical Chemistry, Meyers, R. A., Ed. John Wiley & Sons, Inc.: Chichester, 2000; pp 10815-10837.



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68. Grewer, D. M.; Young, R. F.; Whittal, R. M.; Fedorak, P. M., Naphthenic acids and other acidextractables in water samples from Alberta: What is being measured? Sci. Total. Environ. 2010, 408, (23), 5997-6010. 69. Scott, A. C.; Young, R. F.; Fedorak, P. M., Comparison of GC–MS and FTIR methods for quantifying naphthenic acids in water samples. Chemosphere 2008, 73, (8), 1258-1264. 70. Huang, W. Y.; Brigante, M.; Wu, F.; Mousty, C.; Hanna, K.; Mailhot, G., Assessment of the Fe (III)–EDDS complex in Fenton-like processes: from the radical formation to the degradation of bisphenol A. Environ. Sci. Technol. 2013, 47, (4), 1952-1959. 71. De Laat, J.; Gallard, H., Catalytic decomposition of hydrogen peroxide by Fe (III) in homogeneous aqueous solution: mechanism and kinetic modeling. Environ. Sci. Technol. 1999, 33, (16), 27262732. 72. Gallard, H.; De Laat, J., Kinetic modelling of Fe (III)/H2O2 oxidation reactions in dilute aqueous solution using atrazine as a model organic compound. Water Res. 2000, 34, (12), 3107-3116. 73. Huang, C.; Shi, Y.; El-Din, M. G.; Liu, Y., Treatment of oil sands process-affected water (OSPW) using ozonation combined with integrated fixed-film activated sludge (IFAS). Water Res. 2015, 85, 167-176. 74. Pourrezaei, P.; Drzewicz, P.; Wang, Y.; Gamal El-Din, M.; Perez-Estrada, L. A.; Martin, J. W.; Anderson, J.; Wiseman, S.; Liber, K.; Giesy, J. P., The impact of metallic coagulants on the removal of organic compounds from oil sands process-affected water. Environ. Sci. Technol. 2011, 45, (19), 8452-8459. 75. Garcia-Garcia, E.; Pun, J.; Perez-Estrada, L. A.; Gamal-El Din, M.; Smith, D. W.; Martin, J. W.; Belosevic, M., Commercial naphthenic acids and the organic fraction of oil sands process water downregulate pro-inflammatory gene expression and macrophage antimicrobial responses. Toxicol. Lett. 2011, 203, (1), 62-73. 76. Clayden, J.; Greeves, N.; Warren, S., Organic Chemistry 2012. In Oxford University Press: Oxford, New York: 2012. 77. Pourrezaei, P.; Alpatova, A.; Khosravi, K.; Drzewicz, P.; Chen, Y.; Chelme-Ayala, P.; Gamal El-Din, M., Removal of organic compounds and trace metals from oil sands process-affected water using zero valent iron enhanced by petroleum coke. J. Environ. Manage. 2014, 139, 50-58. 78. Dinh, T. V., Multicomponent analysis by synchronous luminescence spectrometry. Analytical Chemistry 1978, 50, (3), 396-401. 79. Larson, R. A.; Weber, E. J., Reaction mechanisms in environmental organic chemistry. CRC press: Boca Raton, Florida, 1994. 80. Shu, Z.; Li, C.; Belosevic, M.; Bolton, J. R.; El-Din, M. G., Application of a solar UV/chlorine advanced oxidation process to oil sands process-affected water remediation. Environ. Sci. Technol. 2014, 48, (16), 9692-9701. 81. Garrett, R. M.; Pickering, I. J.; Haith, C. E.; Prince, R. C., Photooxidation of crude oils. Environ. Sci. Technol. 1998, 32, (23), 3719-3723. 82. Jacobs, L. E.; Weavers, L. K.; Chin, Y. P., Direct and indirect photolysis of polycyclic aromatic hydrocarbons in nitrate‐rich surface waters. Environmental Toxicology and Chemistry 2008, 27, (8), 1643-1648. 83. von Sonntag, C.; Dowideit, P.; Fang, X.; Mertens, R.; Pan, X.; Schuchmann, M. N.; Schuchmann, H.-P., The fate of peroxyl radicals in aqueous solution. Water Science and Technology 1997, 35, (4), 9-15. 84. Lindsey, M. E.; Tarr, M. A., Inhibition of hydroxyl radical reaction with aromatics by dissolved natural organic matter. Environ. Sci. Technol. 2000, 34, (3), 444-449. 85. Buxton, G. V.; Greenstock, C. L.; Helman, W. P.; Ross, A. B., Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals. Journal of Physical and Chemical Reference Data 1988, 17, 513-886.


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86. Semadeni, M.; Stocker, D. W.; Kerr, J. A., The temperature dependence of the OH radical reactions with some aromatic compounds under simulated tropospheric conditions. Int. J. Chem. Kinet 1995, 27, (3), 287-304. 87. Maki, H.; Sasaki, T.; Harayama, S., Photo-oxidation of biodegraded crude oil and toxicity of the photo-oxidized products. Chemosphere 2001, 44, (5), 1145-1151. 88. Payne, J. R.; Phillips, C. R., Photochemistry of petroleum in water. Environ. Sci. Technol. 1985, 19, (7), 569-579. 89. Bobinger, S.; Andersson, J. T., Photooxidation products of polycyclic aromatic compounds containing sulfur. Environ. Sci. Technol. 2009, 43, (21), 8119-8125. 90. Afzal, A.; Drzewicz, P.; Perez-Estrada, L. A.; Chen, Y.; Martin, J. W.; Gamal El-Din, M., Effect of molecular structure on the relative reactivity of naphthenic acids in the UV/H2O2 advanced oxidation process. Environmental Science & Technology 2012, 46, (19), 10727-34. 91. Frankenfeld, J. W. In In "Proceedings of Joint Conference on Prevention and Control of Oil Spills", Joint Conference on Prevention and Control of Oil Spills, American Petroleum Institute: Washington, D.C., 1973; American Petroleum Institute: Washington, D.C., 1973; pp 485-498. 92. Rubin, M.; Martell, A. E., The implications of trace metal-nitrilotriacetetic acid speciation on its environmental impact and toxicology. Biological trace element research 1980, 2, (1), 1-19. 93. Anderson, R. L.; Bishop, W. E.; Campbell, R. L.; Becking, G. C., A review of the environmental and mammalian toxicology of nitrilotriacetic acid. CRC critical reviews in toxicology 1985, 15, (1), 1102. 94. Levey, S.; Lasichak, A.; Brimi, R.; Orten, J.; Smyth, C. J.; Smith, A. H., A study to determine the toxicity of fumaric acid. Journal of the American Pharmaceutical Association 1946, 35, (10), 298304. 95. Stolzberg, R. J.; Hume, D. N., Rapid formation of iminodiacetate from photochemical degradation of iron (III) nitrilotriacetate solutions. Environ. Sci. Technol. 1975, 9, (7), 654-656. 96. Trott, T.; Henwood, R. W.; Langford, C. H., Sunliight photochemistry of ferric nitrilotriacetate complexes. Environmental Science & Technology 1972, 6, (4), 367-368. 97. Wang, N.; Chelme-Ayala, P.; Perez-Estrada, L.; Garcia-Garcia, E.; Pun, J.; Martin, J. W.; Belosevic, M.; Gamal El-Din, M., Impact of ozonation on naphthenic acids speciation and toxicity of oil sands process-affected water to Vibrio fischeri and mammalian immune system. Environ. Sci. Technol. 2013, 47, (12), 6518-6526. 98. Hu, X.; Wang, X.; Ban, Y.; Ren, B., A comparative study of UV–Fenton, UV–H2O2 and Fenton reaction treatment of landfill leachate. Environ. Technol. 2011, 32, (9), 945-951. 99. Drzewicz, P.; Afzal, A.; Gamal El-Din, M.; Martin, J. W., Degradation of a model naphthenic acid, cyclohexanoic acid, by vacuum UV (172 nm) and UV (254 nm)/H2O2. J. Phys. Chem. A 2010, 114, (45), 12067-12074.

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712 713 714 715

716 717

Figure 1. a) Cyclic voltammograms of Fe(III/II)NTA and Fe(III/II)EDDS at pH 7 and b) half-wave potentials of

718

Fe(III/II)NTA and Fe(III/II)EDDS at pHs 7, 8 and 9 ([Fe]0 = 1mM and [NTA]0 = [EDDS]0 = 2 mM).

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Figure 2. a) NTA degradation in MilliQ (MQ) water and in OSPW and b) EDDS degradation in MQ water ([NTA]0

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= 0.72 mM for UV-NTA, UV-Fe(III)NTA, and UV-NTA-Fenton; [EDDS]0 = 0.72 mM for UV-EDDS, UV-

737

Fe(III)EDDS, and UV-EDDS-Fenton; [Fe]0 = 0.089 mM for UV-Fe(III)NTA/EDDS and UV-NTA/EDDS-Fenton;

738

and [H2O2]0 = 5.88 mM for UV-NTA/EDDS-Fenton).

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754 755

Figure 3. The removal of acid extractable organic fraction (AEF) with increasing H2O2 dose in the UV-

756

NTA/EDDS-Fenton processes ([Fe]0 = 0.089 mM, [NTA]0 = [EDDS]0 = 0.72 mM, and 30 min UV irradiation).

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Figure 4. a) 1H NMR spectra for raw OSPW and the OSPW treated with the UV-NTA/EDDS-Fenton processes

777

([H2O2]0 = 5.88 mM and 4.41 mM for UV-NTA-Fenton and UV-EDDS-Fenton, respectively. [Fe]0 = 0.089 mM and

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[NTA]0 = [EDDS]0 = 0.72 mM, and 30 min UV irradiation) (the values above the peaks are the relative peak area

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with the peak of the internal standard DMSO set as 100), b) SFS of the raw OSPW and the OSPW treated with the

780

UV-NTA/EDDS-Fenton processes ([Fe]0 = 0.089 mM, [NTA]0 = [EDDS]0 = 0.72 mM, and 30 min UV irradiation).

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Figure 5. a) Effect of the H2O2 dose on the acute toxicity of OSPW treated with the UV-NTA/EDDS-Fenton

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processes ([Fe]0 = 0.089 mM, [NTA]0 = [EDDS]0 = 0.72 mM, and 30 min UV irradiation) and b) acute toxicity of

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NTA, EDDS, Fe(III)NTA/EDDS and their products in MilliQ water ([H2O2]0 = 5.88 mM, [Fe]0 = 0.089 mM, and

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[NTA]0 = [EDDS]0 = 0.72 mM). The numbers (0, 2, 4, 8, 12) denote the sampling time.

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799 800

Figure 6. NA distribution in a) raw OSPW, and OSPW treated with b) UV-NTA-Fenton, c) UV-H2O2, and d) NTA-

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Fenton ([H2O2]0 = 5.88 mM for all three processes. [Fe]0 = 0.089 mM and [NTA]0 = 0.72 mM for UV-NTA-Fenton

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and NTA-Fenton. 30 min UV irradiation for UV-NTA-Fenton and UV-H2O2).

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Abstract Art:

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Comparison of Nitrilotriacetic Acid and [S, S]-Ethylenediamine-N, N

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