Oxidative Degradation of Nalidixic Acid by Nano-magnetite via Fe2+

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Oxidative Degradation of Nalidixic Acid by Nanomagnetite via Fe2+/O2-Mediated Reactions Sandy Ardo, Sylvie Nélieu, Georges Ona-Nguema, Ghislaine Delarue, Jessica Brest, Elsa Pironin, and Guillaume Morin Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/es505649d • Publication Date (Web): 10 Mar 2015 Downloaded from http://pubs.acs.org on March 12, 2015

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Environmental Science & Technology

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Oxidative Degradation of Nalidixic Acid by Nano-

2

magnetite via Fe2+/O2-Mediated Reactions

3

SANDY G. ARDO†, SYLVIE NELIEU‡, GEORGES ONA-NGUEMA†, GHISLAINE

4

DELARUE‡, JESSICA BREST†, PIRONIN ELSA†, GUILLAUME MORIN†

5 6 7

†

Institut de Minéralogie, de Physique des Matériaux et de Cosmochimie (IMPMC), UMR

7590, CNRS – UPMC – IRD – MNHN, 4 Place Jussieu F-75252 Paris Cedex 05, France ‡

INRA, UR251 PESSAC, Route de Saint-Cyr F-78026 Versailles Cedex, France

8 9 10 11 12 13 14 15

Corresponding author:

Guillaume Morin;

16

E-mail: [email protected]

17

Tel: +33 1 44 27 75 04

18

Fax: +33 1 44 27 37 85

19

ACS Paragon Plus Environment

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ABSTRACT

21

Organic pollution has become a critical issue worldwide due to the increasing input and

22

persistence of organic compounds in the environment. Iron minerals are potentially able to

23

degrade efficiently organic pollutants sorbed to their surfaces via oxidative or reductive

24

transformation processes. Here, we explored the oxidative capacity of nano-magnetite (Fe3O4)

25

having ~12 nm particle size, to promote heterogeneous Fenton-like reactions for the removal

26

of nalidixic acid (NAL), a recalcitrant quinolone antibacterial agent. Results show that NAL

27

was adsorbed at the surface of magnetite and was efficiently degraded under oxic conditions.

28

Nearly 60% of this organic contaminant was eliminated after 30 min exposure to air bubbling

29

in solution in the presence of an excess of nano-magnetite. X-ray diffraction (XRD) and Fe K-

30

edge X-ray absorption spectroscopy (XANES and EXAFS) showed a partial oxidation of

31

magnetite to maghemite during the reaction, and four by-products of NAL were identified by

32

liquid chromatography-mass spectroscopy (UHPLC-MS/MS). We also provide evidence that

33

hydroxyl radicals (HO•) were involved in the oxidative degradation of NAL, as indicated by

34

the quenching of the degradation reaction in the presence of ethanol. This study points out the

35

promising potentialities of mixed valence iron oxides for the treatment of soils and

36

wastewater contaminated by organic pollutants.

37


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TOC

39 40 Byproducts Nalidixic acid

ROS O2

Maghemite Fe23+O3 Data Fit Residual Magnetite Maghemite

Normalized Absorbance (a.u)

Magnetite Fe2+Fe23+O4

7100

41

7120

7140

7160

7180

7200

Energy (eV)


3


42

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INTRODUCTION

43 44

The quinolone antimicrobials have found wide application in medicine since the late 1960s.

45

The widespread use of this class of antibiotics and their incomplete removal from wastewaters

46

1, 2, 3

47

significant increase in antimicrobial resistance to these molecules causing potential health

48

problem

49

oxo-[1,8]naphthyridine-3-carboxylic acid) is the first synthesized antimicrobial quinolone.

50

NAL is an ionizable, non-biodegradable photosensitive molecule

51

function having a pKa of 5.95

52

occurring in soils and sediments such as Fe- and Al-oxides

53

manufactured nanoparticles such as nano-magnetite

54

(2014)

55

value near the pKa of NAL, while desorption occurred under alkaline conditions, as expected

56

for anionic species. Furthermore, they suggested that NAL forms inner-sphere or hydrogen

57

bonded surface complexes onto magnetite. In addition, biodegradation

58

photodegradation

59

been explored as potential transformation processes. The rate and extent of biological

60

remediation of quinolones is generally slow and the reported degradation levels do not exceed

61

50%

62

biological treatment 19.

63

Fenton-based oxidation techniques such as photo-Fenton

64

reactions (Fe3+/H2O2)

65

increasing attention has been paid to the Fenton-reaction in soil and water treatment because

66

of its high oxidative power towards refractory organic pollutants that are difficult to eliminate

have led to their residual occurrence in natural waters4, soils 5 and sediments 6, and to a

7, 8

14

. Nalidixic acid (NAL) with molecular formula C12H12N2O3 (1-ethyl-7-methyl-4-

11

9, 10

, with a carboxylic acid

. NAL has been shown to efficiently sorb onto minerals

14

12

and clays

13

as well as

. Recently, the study by Usman et al.

showed that NAL adsorption at the surface of magnetite was the highest at a pH

9, 15, 16

17

15,

16

and

of quinolones as well as advanced oxidation processes (AOPs)18 have

. In contrast, ozonation of NAL was shown to be very efficient compared to

21

20

and homogenous Fenton-like

have also been applied to NAL transformation. In the last decades,


4

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22

67

by other methods

68

H2O2, leading to the formation of reactive oxygen species (ROS)

69

radical (HO•) is a short lived, highly reactive and non-selective oxidizing agent. Therefore, it

70

is capable of decomposing a wide variety of organic and inorganic pollutants

71

this reaction is limited to a narrow pH range with an optimum at pH 3, which may limit its in-

72

situ application to natural environments

73

and clays

74

Fenton reactions in the presence of H2O2 over a wider pH range, although being more

75

efficient under acidic conditions. For instance, Matta et al. (2008) 27 reported the oxidation of

76

2,4,6-trinitotoluene at neutral pH, after reaction with various iron-bearing minerals including

77

magnetite (Fe3O4) in the presence of hydrogen peroxide. Magnetite (Mt), especially in nano-

78

crystalline form, has attracted attention in the field of environmental remediation

79

high sorption ability for toxic metals

80

demonstrated. Magnetite nanoparticles present furthermore the advantage of being easily

81

removed from aqueous suspensions using low magnetic fields 41. As well, magnetite has been

82

identified as a strong sorbent for organic pollutants such as chlorotetracycline

83

acid

84

using a strong oxidant such as H2O2, for the degradation of 4,6-dinitro-o-cresol

85

hydrocarbons

86

ibuprofen 48. Beyond the use of strong oxidants that may have adverse effects on ecosystems,

87

a few recent studies have shown the role of dissolved oxygen (O2) as first electron acceptor in

88

magnetite based heterogeneous Fenton processes. Ona-Nguema et al. (2010)

89

rapid arsenic(III) oxidation at Mt surfaces in the presence of dissolved oxygen via Fe2+

90

mediated Fenton-like reactions. Application of such heterogeneous Fenton processes using

91

magnetite and O2 to oxidative treatment of organic molecules is still scarcely documented.

14

31, 32

. The homogeneous Fenton-reaction relies on ferrous iron oxidation by

43

35

. Among them, hydroxyl

. Fe(II)-containing (hydr)oxides

as well as zero valent iron

and ciprofloxacin

45

25

23

33

26-28

24

. However,

, zeolites

29, 30

have been shown to mediate heterogeneous

, metalloids

36-38

and actinides

39, 40

34

, since its

has been widely

42

, nalidixic

and an efficient substrate in the heterogeneous Fenton reaction

, polycyclic aromatic hydrocarbons

46

, nitrophenol


47

44

, oil

, carbamazepine and

49

proved the

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92

Fang et al. (2013) 50 recently reported the degradation of 2-chlorobiphenyl by Mt using O2 as

93

oxidant.

94

Here, we investigated the oxidative degradation of nalidixic acid by nano-magnetite in the

95

presence of dissolved O2 as a function of pH, in the dark and in the absence of strong oxidant.

96

Magnetite oxidation into maghemite led to the degradation of NAL under acidic to weakly

97

alkaline pH conditions; HO• radicals were identified as major oxidizing agent in this process.

98 99 100

MATERIALS AND METHODS

101 102

Nano-magnetite synthesis and nalidixic acid degradation experiments were conducted in a

103

COY glovebox under N2 atmosphere (< 10 ppm of O2) with O2-free de-ionized water (18.2

104

M.cm; Millipore Milli-Q system) degassed by bubbling N2 (Alphagaz 1, Air liquide) for 45

105

min at 80°C.

106 107

Nano-magnetite

and

nano-maghemite

synthesis.

Nano-magnetite

108

synthesized by aqueous coprecipitation of Fe2+ and Fe3+ ions in O2-free deionized water raised

109

at pH 12 by NaOH 49. Solids were then washed three times with O2-free deionized water and

110

vacuum-dried in a glove-box. X-ray diffraction (XRD) indicated that they consisted of pure

111

nanoMt with Mean Coherent Domain size of 11  1 nm. High Resolution Transmission

112

Electron Microscopy indicated cuboctahedral single domain nanoMt particles (Figure SI-1),

113

with average diameter of 13 nm as determined from the observation of ~30 particles in the 9 -

114

18 nm range. Average particle size of ~12 ± 1 nm obtained from both XRD and HRTEM

115

analyses was consistent with the surface area of 93  2 m2 g-1 determined by N2 BET surface

116

analysis (BELSORP-mini, BEL Japan, Inc). Nano-maghemite (NanoMh) was synthesized by


(NanoMt)

was

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117

heating nano-magnetite to 200 °C for 18 h. The powder turned from black to brown. The

118

purity was confirmed by XRD (Figure SI-2).

119 120

Nalidixic acid degradation experiments. Two types of experiments were conducted,

121

allowing or not the pH to evolve over the course of the experiment. In the first type of

122

experiments in which pH was allowed to change, referred to as ‘variable-pH experiments’, the

123

suspensions were rotary shaken and the starting pH value was 6.5 ± 0.2. This value falls

124

within the pH range of maximum sorption for NAL 14 and of environmental pH values. In the

125

second type of experiments referred to as ‘fixed-pH experiments’, the pH was automatically

126

maintained to initial values of 4.5, 6.5 or 8.5, by adding appropriate volumes of HCl (1 M)

127

and NaOH (1 M) solutions using two Titronic® universal Piston Burettes, according to a

128

computer feedback control loop based on pH measurements obtained by a Hanna HI 221 pH-

129

meter. Suspensions were stirred using a propeller mixer.

130

All experiments were performed in darkness in duplicates with an initial NAL concentration

131

of 2.9 mg L-1 (i.e., 12.5 µmol L-1 as NAL molecular weight is 232 g mol-1), with and without

132

nanoMt at a concentration of 15.6 g L-1 (i.e., 1.45 103 m2 L-1)) and at a ionic strength of 10

133

mM NaCl, in a total volume of 32 mL. For this purpose, a volume of 5 mL of a nanoMt stock

134

suspension (100 g L-1), or of water, was added to 27 mL of a NAL and NaCl solution

135

previously adjusted to the desired starting pH. The stock suspensions of magnetite were also

136

previously adjusted for 2 days to the desired starting pH and ultra-sonicated for 45 min prior

137

to use. The pH was readjusted to the desired starting value after mixing. All samples were

138

then maintained under anoxic conditions for 3 days. After this period, the samples (except

139

anoxic controls) were exposed each day to 2 × 15 min air-bubbling (25 mL min-1) with an

140

interval of three hours between the two air-bubbling operations. This treatment was applied

141

for 3 consecutive days from day 3 to day 5. In order to investigate the role of hydroxyl


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142

radicals in our experiments, the variable-pH experiments were repeated in the presence of

143

10% vol. of ethanol, which is used as HO• scavenger. Suspension samples were collected for

144

analysis 24 hours after each of the 2 × 15 min bubbling, i.e. at day 4, 5 and 6. Reaction was

145

quenched immediately after sampling by adding 5 mL of methanol. An aliquot was taken to

146

determine total Fe(II), by dissolution of the solid phase in 5 M HCl and measure using the

147

1,10-phenanthroline colorimetric method at 510 nm 51. Initial dissolved Fe(II) concentrations

148

were measured using the same colorimetric method, after 0.2 µm filtration. The suspension

149

was then centrifuged for 30 min at 8000 rpm. The supernatant was filtered through 0.2 µm

150

Acrodisc® PES filters of 13 mm diameter for further HPLC analyses. In order to recover

151

sorbed NAL and by-products, the solid was then washed with 20 mL O2-free water at pH 12

152

and centrifuged three times for 30 min. Supernatants of these desorption treatments were

153

pooled and filtered for further HPLC analyses. Efficiency of this desorption treatment was

154

previously optimized for NAL as a function of the desorbing solution pH (Figure SI-3). Solids

155

were vacuum-dried and stored in anoxic containers for further XRD and XAS analyses.

156 157

High

performance

liquid

chromatography

(HPLC).

Aqueous

and

desorbed

158

concentration of NAL was measured by HPLC using a Dionex System including an ASI100T

159

autosampler, a P580 pump, a STH585 column oven and a UVD380S UV-photodiode array

160

detector. Separation was done on a 125 mm × 2 mm 3 µm Nucleodur C18 HTec reversed

161

phase column (Macherey-Nagel®, Düren, Germany) at 20°C, using a mixture of 25%

162

acetonitrile and 75% water acidified by 0.1% acetic acid, at 0.33 mL/min flow rate. Samples

163

were kept at 15°C and protected from light in the autosampler before being injected. The

164

injection volume was 5 µL. Quantifications were carried out at 254 nm using external

165

calibration. Desorbed concentration of NAL was obtained as follows: Cdes = Caq.Vdes / Vi ,

166

where Cdes is the desorbed concentration of NAL (mg L-1), Caq is the NAL concentration


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167

measured in the desorbing water (mg L-1), Vdes is the volume of the desorbing water (mL) and

168

Vi is the initial volume of the batch (mL).

169 170

Ultra-high-performance liquid chromatography - tandem mass spectrometry

171

(UHPLC-MS/MS). The identification of NAL by-products was performed by, using a Waters

172

Acquity UPLC system coupled to a triple-quadrupole mass spectrometer (TQD, Waters,

173

France) via an electrospray ionization interface. Chromatographic separation was achieved on

174

an UHPLC C18 column, with an acetonitrile/water (both acidified) gradient. The ionization

175

was performed in positive and negative modes. Analyses in MS-MS were performed in

176

daughter mode using argon for collision-induced dissociation. Prior to injection, samples

177

(acidified to pH  3) were concentrated through solid phase extraction (SPE) on Oasis HLB

178

120 mg cartridges (Waters). Detailed analytical procedures are given in Supporting

179

Information.

180 181

X-ray Diffraction (XRD). Powder patterns were collected using a Panalytical X’Pert Pro

182

MPD diffractometer equipped with an anoxic sample chamber to avoid any oxidation during

183

the measurements. Data were collected using Co K radiation (=0.17889 nm) in continuous

184

scan mode within 10-80° 2 range with a step size of 0.0167°.

185 186

X-ray Absorption Spectroscopy (XAS). Data were recorded at the iron K-edge at 40K on

187

beamline BM23 at the European Synchrotron Radiation Facility (ESRF) in transmission

188

detection mode, averaging two scans for each sample. Energy was calibrated by setting to

189

7112 eV, the first inflection point of the iron K-edge for a Fe(0) foil recorded in double

190

transmission setup. X-ray absorption near edge structure (XANES) and Extended X-ray

191

absorption fine structure (EXAFS) spectra were normalized and background subtracted using


9


52

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. The oxidation state and local structural environment of iron were

192

the ATHENA code

193

determined by linear least-squares fitting of XANES and EXAFS spectra, using linear

194

combinations of experimental spectra of synthetic nano-magnetite and nano-maghemite

195

model compounds.

196 197

RESULTS AND DISCUSSION

198 199

Variable pH experiments. Figure 1 displays the residual fraction of NAL measured in

200

batch experiments conducted with and without nanoMt at starting pH 6.5, and in which pH

201

was allowed to change over the course of the experiment. Dissolved and sorbed fractions

202

corresponding to these residual values are reported in Table SI-1 and the evolution of the pH

203

in these experiments is displayed in Figure SI-4. These results indicate that, in the oxic and

204

anoxic controls without nanoMt, NAL remained in the dissolved phase and its concentrations

205

remained constant. Addition of nanoMt led to rapid sorption of 97-99% NAL from solution as

206

indicated by the decrease in dissolved concentration of NAL (Table SI-1). Therefore,

207

evaluating the degradation of NAL required to measure the quantity of NAL recovered after

208

desorption from the solid phase (Table SI-1). The residual NAL proportion (Figure 1) was

209

thus determined by adding this desorbed quantity of NAL to the low quantity of residual

210

dissolved NAL (Table SI-1). Under anoxic conditions, no degradation was observed with

211

residual NAL ranging between 80 to 96% of the initial NAL concentration (2.9 mg L-1). This

212

slight decrease in residual NAL under anoxic conditions is due to incomplete NAL desorption

213

yield (Figure SI-3). In contrast, under oxic conditions a fast initial decrease of NAL

214

concentration was observed at day 4 (after 2 × 15 min of air bubbling) leaving only 36% of

215

residual NAL. It is then followed by a slower degradation at day 5 and 6 (after 4 × 15 min and

216

6 × 15 min air-bubbling) (Figure 1; Table SI-1). In the presence of ethanol as HO•


10

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217

scavenger50, the efficiency of NAL degradation decreased by a factor of about two as shown

218

in Figure 1b and Table SI-2. This result suggests that HO• radicals play an important role in

219

the degradation of NAL in our experiments and that other reactive oxygen species may likely

220

be involved.

221 222

Solid phase evolution. XRD analyses of the solids collected over the course of the

223

variable-pH experiments conducted under oxic conditions (Figure SI-2) indicate a partial

224

oxidation of nanoMt to nanoMh after air exposure. Indeed the shift in the (511) and (440)

225

Bragg reflections toward higher diffraction angle values indicates a decrease in unit cell

226

dimensions, in agreement with the smaller unit cell of Mh (8.340 Å) 53 compared to Mt (8.390

227

Å)

228

analyses that indicate a progressive oxidation of nanoMt with increasing exposure time to air

229

in the oxic experiments (Table SI-3). The corresponding decrease in the quantity of total

230

Fe(II) is in good agreement with that determined by colorimetric measurement (Figure 2b and

231

Table SI-1). The absence of change in the width of the XRD Bragg reflections suggests that

232

this partial oxidation process has proceeded without significant change in particle size. No

233

other iron phase was detected by XRD, suggesting that other iron oxidation products that

234

could have also formed as separate phases or as coating on the nanoMt surface would not be

235

crystalline. Furthermore, XANES and EXAFS linear combination fitting indicates that Mh is

236

the major ferric mineral phase observed, without evidence for the presence of poorly

237

crystalline phases, such as ferrihydrite (Fh). Indeed attempts to fit the reacted solids spectra

238

with 2lines Fh as ferric iron fitting component increased the fitting residue and R-factor

239

values by a factor of 3 (Figure SI-6). Moreover fitting the data including both Mh and 2Line

240

Fh as fitting components yields Fh proportions below 3% of total iron, which is below the

241

detection limit of the method 55. These mineralogical results indicate a partial oxidation of the

54

. This observation is fully supported by XANES (Figure 2a) and EXAFS (Figure SI-5)


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242

nanoMt particles without major structural change,as previously observed for magnetite to

243

maghemite transformation in aqueous suspension56-58. Finally, adding ethanol did not affect

244

the oxidation rate of nanoMt (Figure SI-7).

245 246

Evolution of the pH. In the variable-pH experiments reported above, in which the pH

247

was initially set at 6.5 and was not maintained constant over the course of the experiment, a

248

significant decrease of the pH from 6.5-7.0 to 4.0 was observed in the oxic samples even in

249

the absence of NAL (Figure SI-4 and Table SI-1). The overall equation of the oxidation of Mt

250

into Mh is not pH dependant:

251 252 253

Fe3O4 (s) +

1 3 O2 (aq)  Fe2O3 4 2

(1)

254

However, the observed decrease of pH is related to the oxidation of aqueous Fe2+ that was in

255

equilibrium with magnetite at pH 6.5 at the start of the experiment (Figure SI-8). Indeed,

256

adjustment of the magnetite suspension to pH 6.5 at the start of the experiment required acid

257

addition (See Materials and Methods) and thus lead to incongruent dissolution of Mt that can

258

be written as follows at near neutral pH:

259

Fe3O4(s) + 2H+  Fe2+(aq) + Fe2O3 (s) + H2O

(2)

260

According to Gibbs free energy of formation reported for Mt, Mh and aqueous Fe2+ by

261

Beverskog and Puigdomenech, (1996)

262

start of the pH 6.5 experiments is 2.5 10-5 M (1.4 mg L-1), which is ten times lower than the

263

value actually measured by colorimetric analysis (10.3 mg L-1; Table SI-4). This discrepancy

264

could be related to lower stability of nanosized crystals compared to their bulk counterpart,

265

due to interfacial excess free energy 60. Since Mh is the sole ferric phase observed at the end

266

of our experiments, oxidation of aqueous Fe2+ and precipitation of Fe(III) as Mh explains the

267

release of protons responsible for the pH decrease in our system, as follows:

59

the expected concentration of aqueous Fe2+ at the


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268 269 270 271


Fe2+(aq) +

1 1 O2 + H2O  Fe2O3 + 2H+ 2 4

(3)

In order to properly evaluate the effect of pH on the oxidative degradation of NAL by   nanoMt, we conducted fixed-pH experiments.

272 273

Fixed-pH experiments. Fenton reactions are well known to be pH-dependent since pH

274

affects the generation of ROS 61. Accordingly, oxidative degradation of organic pollutants by

275

magnetite-induced Fenton like reactions has been shown to be pH dependent, both in the

276

presence of strong oxidants

277

pH over the course of the experiments was not reported. In the present study, we investigated

278

the efficiency of NAL degradation in fixed pH experiments in the presence of O2. The

279

degradation efficiency of NAL at fixed pH 4.5 was higher than those at pH 6.5 and 8.5

280

(Figure 3; Table SI-4), which is consistent with the pH trend observed by Fang et al. (2013) 50

281

for the degradation of 2-chlorobiphenyl. Moreover, we demonstrate that magnetite can

282

degrade NAL in the presence of O2 at constant near neutral (pH 6.5) and even alkaline (pH

283

8.5) conditions (Figure 3a). The proportion of total initial Fe(II) oxidized during the reactions

284

did not significantly vary with pH. It ranged between 28 and 39 % (Figure 3b; Table SI-4),

285

which was similar to the 30% value observed in the variable experiment corresponding to the

286

sample 2 × 15 min air bubbling (Table SI-1). The concentration of initial dissolved Fe(II) in

287

equilibrium with nanoMt significantly increased with decreasing the initial pH value (Figure

288

SI-8 ; Table SI-4) in agreement with incongruent dissolution of magnetite as written in

289

Eq. (2). After exposure to oxygen, the concentrations of dissolved Fe(II) is very low (Figure

290

SI-8; Table SI-4) in all pH conditions investigated (4.5 – 8.5) indicating total oxidation of

291

aqueous Fe2+, likely to Mh (Eq. 3), as also observed in the variable pH experiments.

62

or of O2

50

. However, in this latter study, the evolution of the

292


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293

Identification of NAL by-products. In variable pH experiments, ~80% of the initial NAL

294

was removed after exposure to 6 × 15 min of air bubbling. In order to identify the main by-

295

products generated during the oxidative degradation of NAL, samples were pre-concentrated

296

by SPE and analysed by UHPLC-MS(-MS) in negative and positive ionization modes. After

297

reaction with oxygen, LC-MS revealed 5 major transformation products in the desorption

298

water that are referred to as I to V (Figure SI-9). Molecular masses of these transformation

299

products were determined in MS positive and negative modes, according to pseudo-molecular

300

ions [MH]+, and/or either [M + Na]+ adduct or [M - H]-. As compared to NAL, they present

301

mass changes consistent with a single modification (N-deethylation because of -28 Da shift

302

for III, hydroxylation because of +16 Da shift for IV and V) or a multi-step one (-24 and -22

303

Da shifts for I and II, respectively).

304

The MS/MS spectra obtained by collision induced dissociation of [MH] + ions (Figure 4)

305

show that nalidixic acid and its by-products present common fragmentations. For instance, all

306

except the compound I present as first fragmentation step a loss of H2O (-18 Da), according to

307

a known mechanism from carboxylic group 63, and/or from an hydroxyl function (compounds

308

IV and V). Further fragmentation includes in all cases except II a loss of 44 Da (CO2), from

309

the initial carboxylic group or through an oxetone fragment intermediate. Similarly, several

310

successive losses of 28 Da were observed for NAL, compounds II, III and V, which could

311

correspond to the formation of C2H4 or CO. The first 28 Da loss from NAL (m/z 215 to 187

312

transition) was previously demonstrated to correspond to a C2H4 by high resolution mass

313

spectrometry

314

(Figure 4), it confirms that III corresponds to an N-deethylated by-product. On the contrary,

315

the absence of 28 Da loss from compound IV (at least prior to the -44 Da one) allows

316

suspecting that hydroxylation takes place on the pyridone ring. For its isomer V, all

317

successive fragmentations are comparable to that observed for NAL. Position of the hydroxyl

64

. As further fragmentation of the ion m/z 187 is equivalent for NAL and III


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318

on the ethyl group was deduced from negative MS/MS spectrum, presenting successive losses

319

of 44 Da (CO2), 29 Da (H•CO) and 15 Da (•CH3). A third hydroxylated isomer, maybe on the

320

methyl group, was observed at trace level. Compounds III-V were previously observed when

321

NAL was degraded by various AOPs 20, 21, 65. Compounds I and II were identified in the same

322

studies and our MS/MS data are consistent with the formation of these byproducts in the

323

present study. They result from pyridone ring opening, but the proposed pathways differ

324

between the two authors. According to Fan et al. (2012)21, the ring opening results from a HO•

325

attack to C3 position, which presents the higher electron density in NAL structure. As these

326

authors, we did not observe the first intermediate byproduct, but we suspect the presence of a

327

further one, since we observed traces of a compound having a [MH]+ at m/z 223. Further

328

attacks of HO• radicals then induce oxidation to form the observed transformation products II

329

and then I (Figure 5). Concomitantly and according to Sirtori et al. (2009)65, the ring-

330

hydroxylated compound IV could, through successive radical attacks, also lead to the

331

compound I. The hydroxyl radical involvement is ascertained by the reduced NAL

332

degradation observed in the presence of ethanol, acting as scavenger.

333 334 335

Reaction Mechanism. The present study demonstrates that in the absence of oxygen,

336

NAL is only adsorbed at the surface of nanoMt and no degradation is observed. These results

337

are in agreement with previous studies of organic compounds sorption on nanoMt such as

338

chlorotetracycline

339

14

42

and ciprofloxacin

43

and the sorption of NAL onto sediments and Mt

12,

. Similarly, Fe(II) within the Mt structure is not oxidized in the absence of oxygen.

340

Under oxic conditions, NAL is efficiently removed from the aqueous and solid phases in

341

parallel to the oxidation of nanoMt (~ 64% of NAL is eliminated after 30 min of air bubbling

342

and 80% after 90 min). Partial quenching of this degradation process in the presence of


15


Page 16 of 30

343

ethanol as HO• scavenger supports the hypothesis of Fenton-type reactions where Fe(II)

344

oxidation by molecular oxygen leads to the formation of hydroxyl radicals and other highly

345

oxidant reactive oxygen species

346

process involving magnetite and O2 was reported by Ona-Nguema et al., (2010)

347

showed rapid oxidation of As(III) to As(V) upon sorption onto Mt under oxic conditions at

348

neutral pH, as well as, more recently, by Fang et al. (2013)

349

chlorobiphenyl by ROS generated via the oxidation of Mt in presence of O2. Nevertheless,

350

when the oxidative degradation process was demonstrated

351

provided to determine whether the degradation of the contaminant occurred in solution, at the

352

surface of nanoMt or both, and whether the oxidation process involved aqueous Fe2+ ions

353

released from the mineral structure or structural Fe2+ ions at the mineral surface. Although the

354

present study does not neither allow to conclude on these reaction pathways, our results are

355

consistent with oxidative degradation of adsorbed NAL by ROS produced via oxidation of

356

both dissolved and structural Fe2+ ions. Indeed, both NAL and its degradation products were

357

adsorbed to nanoMt (97-99% for NAL), which supports the oxidation of adsorbed NAL by

358

ROS. In addition, oxidation of aqueous Fe2+ was observed in all our experiments conducted

359

at pH ≤ 6.5, which suggest that ROS could have been partly generated via oxidation of

360

dissolved ferrous iron. Aqueous Fe2+ in our experiments mainly originated from incongruent

361

dissolution of magnetite (Eq. 2) after initial equilibration of the magnetite suspension at the

362

initial pH. The mechanism of aqueous oxidation of Fe2+ by dioxygen is well known and can

363

be written as follows, where Eq. (4) is generally considered as the kinetically limiting step,

364

with half-lives of Fe2+ being approximately 45 h, 4.5 h and 27 min at pH 6, 6.5 and 7

365

respectively 66:

61

that could readily oxidize NAL. Similar Fenton-type

50

49, 50

366

Fe2+(aq) + O2



Fe3+ + O2•-

(4)

367

Fe2+(aq) + O2•- + 2H+



Fe3+ + H2O2

(5)


49

who

for the degradation of 2-

no definitive evidence was

16

Page 17 of 30


368

Fe2+(aq) + H2O2 + H+



Fe3+ + HO• + H2O

(6)

369

Fe2+(aq) + HO• + H+



Fe3+ + H2O

(7)

370

According to this pathway, the generation of HO• is facilitated under acidic conditions, since

371

Eq. 5-7 consume protons. Hence, HO• being more oxidant than other ROS species

372

Fenton-related degradation of organics is classically expected to be favoured under acidic

373

conditions, as observed in our experiments and in previous studies 50. Furthermore, acidic pH

374

favors Mt dissolution (Eq. 2) and is thus expected to enhance the importance of dissolved Fe2+

375

as source of ROS in the degradation process.

376

Under alkaline pH conditions, O2•- has been proposed as being the main oxidant

377

superoxide ion O2•- is capable of oxidizing a wide range of organic contaminants but remains

378

a weaker oxidant than HO•, which generally explains the decrease of degradation rate with

379

increasing pH. Besides, a number of recent studies also suggested that ferryl species (Fe(IV))

380

are formed in the Fenton reaction. Although such mechanism is still a matter of debate 66, 68-70,

381

Hug and Leupin (2003) 61 proposed Fe(IV) as the main oxidant of arsenite in the presence of

382

Fe(II) and O2 under alkaline pH, and discounted the role of O2•-.

383

In the heterogeneous Fenton-like process investigated in the present study, we show that

384

magnetite induces the oxidative degradation of NAL at a fixed pH of 8.5, with a concentration

385

of dissolved Fe2+ below the 1,10-phenanthroline colorimetric method quantification limit

386

(0.02 mg L-1). This result suggests that ROS could also be partly generated via oxidation of

387

structural Fe2+ at the surface of the Mt particles, which could be written as follows:

1 H2O 2

388

Fe3O4 (s) + O2 +

389

Fe3O4 (s) + O2•- + H+ +

390

Fe3O4 (s) + H2O2



3 Fe2O3 (s) + O2•- + H+ 2

3 1 Fe2O3 (s) + HO• + H2O 2 2


,

. The

(4’)

1 3 H2O  Fe2O3 (s) + H2O2 2 2



50, 67

50, 61

(5’)

(6’)

17


391

Fe3O4 (s) + HO•



3 1 Fe2O3 (s) + H2O 2 2

Page 18 of 30

(7’)

392 393

It is noteworthy that the global proton-balance of such ROS production process involving

394

structural magnetite-Fe2+ at the magnetite particle surface (Eqs. 4’ to 7’) would be zero. It

395

would thus be favoured over the aqueous Fe2+ pathway (Eqs. 4 to 7) under alkaline

396

conditions, and could also participate to ROS production under acidic conditions.

397

Finally, we observed that the efficiency of the reaction decreased with time in our

398

experiments (Figure 1), leading to maximum oxidation of only 40% of the initial total Fe(II)

399

and degradation of 80% of the initial NAL after 6 × 15 min air bubbling. Such passivation

400

has been reported in several previous studies involving pollutant transformation based on

401

oxidation of nano-magnetite

402

into maghemite could limit formation of ROS because of Fe2+ depletion at the particle

403

surface. Such hypothesis is supported by previous observation of an oxidized layer at the

404

magnetite particle surface after hydrothermal transformation71. Further investigations of the

405

nano-magnetite oxidation mechanism in aqueous systems at room temperature could help

406

optimizing the efficiency of pollutant degradation processes based on nano-magnetite.

49,50,58

. Progressive oxidation of the magnetite particle surface

407 408

Environmental Implications. The present study demonstrates that magnetite efficiently

409

removes nalidixic acid from water involving an adsorption and an oxidation step via a Fenton-

410

type reaction where magnetite is used as a Fe(II) source. Fe(II) oxidation by oxygen leads to

411

the formation of reactive oxygen species able to degrade NAL. Before the oxidation step,

412

NAL was completely adsorbed to the surface of nano-magnetite which suggests that the

413

oxidative degradation reactions mainly occurred at the surface of the magnetite nanoparticles.

414

This study presents a promising ecological process for the removal of organic contaminants


18

Page 19 of 30


415

using mixed iron oxides and dissolved oxygen, since it is applicable without the addition of

416

strong oxidants that may have adverse effects toward ecosystems.

417 418

ASSOCIATED CONTENT.

419

Supporting information. This additional file includes information on the chemicals used

420

for the experiments, on UHPLC-MS/MS analysis conditions, on nano-magnetite

421

characterization, and on NAL desorption protocol. Additional data on the NAL degradation

422

experiments are also reported. This material is available free of charge via the internet at

423

http://pubs.acs.org.

424 425 426

AUTHOR INFORMATION

427

Corresponding Author

428

Email: [email protected]

429 430 431

ACKNOWLEDGEMENTS

432

The authors are indebted to Nicolas Menguy (IMPMC) for his help in collecting the

433

HRTEM images of the nano-magnetite samples, as well as to Olivier Mathon and Manuel

434

Munoz for their technical support during XAS measurements on the BM23 ESRF beamline

435

(Grenoble, France). We also thank the IMPMC Project group for conception and maintenance

436

of the pH regulation system in the glove box (Marc Morand, Frédéric Gélébart) and of the

437

anoxic XRD chamber (Frédéric Gélébart, Ludovic Delbes, Benoit Baptiste). This study was

438

supported by DIM R2DS IdF “Innovations Technologiques” Grant# 2011-11 and by

439

SESAME IdF Grant #1775 for XRD, and Région IdF for UHPLC-MS-MS.

440 441


19


Page 20 of 30

442 443 444

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67. Pham, A. N.; Waite, T. D., Oxygenation of Fe(II) in natural waters revisited: Kinetic modeling approaches, rate constant estimation and the importance of various reaction pathways. Geochim. Cosmochim. Acta 2008, 72, (15), 3616-3630.

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68. Lee, C.; Sedlak, D. L., Enhanced Formation of Oxidants from Bimetallic Nickel-Iron Nanoparticles in the Presence of Oxygen. Environ. Sci. Technol. 2008, 42, (22), 8528-8533.

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69. Pang, S.-Y.; Jiang, J.; Ma, J., Response to Comment on "Oxidation of Sulfoxides and Arsenic(III) in Corrosion of Nanoscale Zero Valent Iron by Oxygen: Evidence against Ferryl Ions (Fe(IV)) as Active Intermediates in Fenton Reaction". Environ. Sci. Technol. 2011, 45, (7), 3179-3180.

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70. Pang, S.-Y.; Jiang, J.; Ma, J., Oxidation of Sulfoxides and Arsenic(III) in Corrosion of Nanoscale Zero Valent Iron by Oxygen: Evidence against Ferryl Ions (Fe(IV)) as Active Intermediates in Fenton Reaction. Environ. Sci. Technol. 2011, 45, (1), 307-312. 71. Chen, S.Y.; Gloter, A.; Zobelli, A.; Wang, L.; Chen, C.H.; Colliex, C., Electron energy loss spectrsocopy and ab initio investigation of iron oxide nanomatreials grown by a hydrothermal process. Phys. Rev. B 2009, 79, (104103), 1-10

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Figure 1. Results of the variable pH experiments conducted at a starting pH of 6.5 in a)

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absence or b) presence ethanol. Total residual NAL fraction including the aqueous and

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desorbed fractions are plotted as a function of time, under oxic (Oxic) or anoxic (Anoxic)

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conditions, in the presence (+ Mt) or absence (w/o Mt) of nanoMt. Each arrow corresponds to

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15 min air bubbling. Error bars correspond to standard deviation calculated from three and

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two independent experiments carried out without and with ethanol, respectively.

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671 672 673 674 Data Fit 1

Residual Magnetite

(a)

Maghemite

XANES measurements Colorimetric measurements

6x15min air bubbling

34% Mh

2x15min air bubbling 32% Mh

Anoxic

Oxidative Degradation of Nalidixic Acid by Nano-magnetite via Fe2+

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