Photodegradation of Veterinary Ionophore Antibiotics under UV and

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Photodegradation of Veterinary Ionophore Antibiotics under UV and Solar Irradiation Peizhe Sun, Spyros G. Pavlostathis, and Ching-Hua Huang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/es5034525 • Publication Date (Web): 24 Oct 2014 Downloaded from http://pubs.acs.org on October 28, 2014

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Environmental Science & Technology

1

Photodegradation of Veterinary Ionophore Antibiotics

2

under UV and Solar Irradiation

3 4

Peizhe Sun, Spyros G. Pavlostathis, Ching-Hua Huang*

5 6

School of Civil and Environmental Engineering, Georgia Institute of Technology, Atlanta,

7

Georgia 30332, United States

8 9

* Corresponding Author. Phone: 404-894-7694; Fax: 404-358-7087.

10

E-mail: [email protected]

11 12

Revised manuscript submitted to

13


14 15

October 24, 2014

16

ACS Paragon Plus Environment


17

ABSTRACT

18

The veterinary ionophore antibiotics (IPAs) are extensively used as coccidiostats and

19

growth promoters, and are released to the environment via land application of animal

20

waste. Due to their propensity to be transported with runoff, IPAs likely end up in surface

21

waters where they are subject to photodegradation. This study is among the first to

22

investigate the photodegradation of three commonly used IPAs, monensin (MON),

23

salinomycin (SAL) and narasin (NAR), under UV and solar irradiation. Results showed

24

that MON was persistent in deionized (DI) water matrix when exposed to UV and

25

sunlight, whereas SAL and NAR could undergo direct photolysis with a high quantum

26

yield. Water components including nitrate and dissolved organic matter had a great

27

impact on the photodegradation of IPAs. A pseudo-steady state kinetic model was

28

successfully applied to predict IPAs’ photodegradation rates in real water matrices.

29

Applying LC/MS/MS, multiple photolytic transformation products of IPAs were

30

observed and their structures proposed. The direct photolysis of SAL and NAR occurred

31

via cleavage on the ketone moiety and self-sensitized photolysis. With the presence of

32

nitrate, MON was primarily degraded by hydroxyl radicals, whereas SAL showed

33

reactivity toward both hydroxyl and nitrogen-dioxide radicals. Additionally, toxicity tests

34

showed that photodegradation of SAL eliminated its antibiotic properties against Bacillus

35

subtilis.


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36


INTRODUCTION

37

Ionophore antibiotics (IPAs) are among the most common veterinary pharmaceuticals

38

used to prevent coccidiosis and promote digestion efficiency for meat production.1,2 In

39

2011, the annual sales of IPAs in the U.S. reached over 4.1 million kg, which made them

40

the second top-selling antimicrobial group next to tetracyclines.3 Among IPAs, monensin

41

(MON), salinomycin (SAL) and narasin (NAR) (Scheme 1) are most commonly used in

42

the livestock industry. IPAs are polyether carboxylic acids which inhibit the growth of

43

coccidia, and have shown adverse effects on Gram-positive bacteria, algae, and

44

protozoa.4-10

45

IPAs, administered in livestock feed, are mostly excreted due to poor absorption or

46

limited degradation in the animals’ digestion systems,11 and consequently are released to

47

the environment via land application of animal manure. Although IPAs are hydrophobic

48

compounds, significant amounts of IPAs can be transported with rainfall runoff from

49

manure-fertilized fields,12,13 which eventually end up in receiving waters. The residual

50

concentrations of IPAs in surface water and water treatment plants have been reported.14-

51

19

52

µg·L-1 in river water near agricultural fields.19 Watkinson et al. examined the occurrence

53

of antibiotics in different environmental water systems, among which MON and SAL

54

were detected at 94% and 21% of all collected samples (n > 84) in surface waters and

55

occasionally in wastewater treatment plant effluents.17

For example, IPAs were found at up to 0.22 µg·L-1 in surface water,16 and at 0.03−0.06

56

Photolysis is known to affect the fate of various pharmaceutical compounds in the

57

aquatic environment. In surface or waste waters, degradation of IPAs may occur due to

58

exposure to solar (in natural water systems) or UV (in treatment facilities) radiation and



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such photodegradation may play an important role in affecting the overall environmental

60

fate of IPAs. To date, however, few studies have evaluated the susceptibility of IPAs to

61

photo-induced degradation. In comparison, the degradation of IPAs via other

62

transformation

63

biodegradation of IPAs was reported in several studies with half-lives of 3−5 days in

64

soil.20-23 IPA’s instability to acid-catalyzed hydrolysis was investigated by our recent

65

study which elucidated transformation mechanisms and products.24 The degradation of

66

IPAs by UV/H2O2 advanced oxidation process (AOP) was also examined by us

67

recently.25 Although studies on the photodegradation of IPAs are scarce, it is expected

68

that some IPAs (e.g., SAL) may undergo direct photolysis due to the presence of the

69

carbonyl moiety in their structures.26

mechanisms

has

been

investigated

previously.

For

example,

70

To address the lack of information in the photo-behaviors of IPAs, the objectives of

71

this study were to investigate the direct and indirect photodegradation of IPAs under

72

environmental conditions, measure the quantum yields, elucidate the transformation

73

products and mechanisms, and evaluate the toxicity of photo-transformation products. 5

R

HO

1

4 2

3

A O

OH

7 8

15

37

32

6

O

9

O

10

11

14 12

16

B 17

38 O

C 21

13 O 18

30

74 75 76 77

33

E

D 24

41 28 40

25 O 29

O

20 OH

OH

27

39

42

19

36

31

Monensin (MON)

35

34

23 26

22

Salinomycin (SAL) (R = -H) Narasin (NAR) (R = -CH3)

Scheme 1. Structures of MON, SAL and NAR.

78 79

MATERIALS AND METHODS

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Chemicals. Suwannee River humic acid (SRHA) was purchased from the

81

International Humic Substances Society (IHSS, St. Paul, MN, USA). Another type of


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humic acid (MPHA) was purchased from MP Biomedicals (Solon, OH, USA). Sources of

83

other chemicals and reagents are provided in the Supporting Information (SI) Text S1. To

84

prepare IPA stock solutions, individual IPA powder was weighted and pre-dissolved in

85

methanol. Then, aliquots of methanolic IPA were evaporated to dryness under vacuum

86

and re-dissolved into DI water at 5 mg·L-1, which was then stored at 4-5°C prior to use.

87

Experiments confirmed that the above sample preparation method versus directly

88

dissolving IPA in DI water yielded the same photolysis rates of IPAs, confirming that

89

residual methanol, if any, was not significant (SI Text S2 and Figure S1). Other stock

90

solution preparation and sample collection are described in SI Text S1.

91 92

Photolysis experimental setup. Sample preparation. Reaction solutions were

93

prepared with 0.5−2.5 mg·L-1 IPA (higher concentrations were used for product

94

generation) in different matrices: DI water, wastewater treatment plant (WWTP)

95

secondary effluent, simulated rainfall runoff from poultry litter (PL)-fertilized land, or

96

simulated PL leachate (matrix characteristics summarized in Table 1). The DI water

97

matrix was maintained at pH 7.0 using 5 mM phosphate buffer, whereas the other water

98

matrices were not adjusted for pH. In the experiments to investigate the effects of nitrate

99

and dissolved organic matter (DOM), aliquots of nitrate or DOM stock solutions were

100

spiked into the DI water matrix to achieve the target concentrations.

101

UV Photolysis. Photolysis of IPAs with UV radiation was studied using a similar

102

setup as previously described,25 and shown in SI Figure S2. Experiments were conducted

103

in a magnetically stirred 100-mL cylindrical quartz reactor kept in a photo-chamber

104

equipped with a 4-W low pressure (LP) UV lamp (G4T5 Hg lamp, Philips TUV4W)



105

peaking at 254 nm at ambient temperature (22oC). Reaction was initiated by exposing the

106

solution to UV irradiation. A sample aliquot was taken at each time interval and stored in

107

a 2-mL amber glass vial prior to LC/MS analysis.

108

Simulated and natural sunlight photolysis. Simulated sunlight was generated by a

109

300-W Xenon lamp (PerkinElmer, PE300BF, full spectral output is shown in SI Figure

110

S3). Reaction solutions were kept in 10-mL quartz tubes, which were held perpendicular

111

to the light source (illustrated in SI Figure S2). For natural sunlight experiments, the

112

quartz tubes were kept vertically on the roof of a laboratory building in Atlanta, GA,

113

USA (latitude 33.8o N). Sunlight photolysis experiments were conducted in April, 2013.

114

Chemical and instrumental analysis. IPAs were analyzed by an Agilent 1100 Series

115

high performance liquid chromatography mass spectrometry (HPLC/MS) system (Agilent,

116

Palo Alto, CA) with a reversed-phase Ascentis RP-amide column (2.1 × 150 mm, 3 µm).

117

The mobile phase contained (A) 0.1% formic acid containing 25% acetonitrile, and (B) a

118

mixture of 50/50 acetonitrile/methanol (v/v). The analytical methods were described in

119

detail previously.27 For elucidation of photo-transformation products, a LC/MS/MS unit

120

(Agilent 1260 Infinity LC system, 6410 Triple Quad MSD, Agilent, Palo Alto, CA) was

121

employed. The LC conditions were similar to those of the HPLC/MS system described

122

above. The MS parameters were set up with fragmentation voltage of 135 V, and

123

collision-induced-dissociation (CID) energy of 50 eV. Other details on instrumental

124

analysis are described in SI Text S3.

125

Toxicity tests for transformation products. The growth of Bacillus subtilis (ATCC

126

6633) was shown to be sensitive to IPA concentration and thus this microorganism was


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127

chosen for the toxicity tests.10,24 The growth inhibition is expressed in Equation 1, in

128

which OD is the optical density. Details of toxicity tests are described in SI Text S4.

129

 OD 600 (Sample)   × 100 Growth Inhibition (%) = 1 −  OD 600 (Control) 

(1)

130 131

RESULTS AND DISCUSSION

132

The photodegradation of MON and SAL was investigated extensively under LP UV

133

(UVC), simulated sunlight and natural sunlight irradiation. Because NAR closely

134

resembles SAL in structure with only one methyl group difference on the A ring (Scheme

135

1), NAR behaved nearly identically to SAL in acid-catalyzed hydrolysis24 and was

136

expected to behave similarly to SAL in photolysis as well. Thus, NAR was investigated

137

only in selected cases.

138 139

Direct photolysis of IPAs under LP UV, simulated sunlight and natural sunlight.

140

Photolysis of IPAs due to direct absorption of light was investigated in DI water buffered

141

by 5 mM phosphate at pH 7. In all dark controls, no significant loss of IPAs was

142

observed for over one week (data not shown). MON was resistant to photodegradation in

143

DI water matrix under exposure to any light source. In contrast, the degradation of SAL

144

was observed, at the fastest rate under UV radiation, followed by simulated sunlight and

145

then natural sunlight (Table 2, SI Figure S4).

146

MON consists of mostly sigma bonds (C−H, C−C, C−O and O−H) except for a

147

carboxylic group at one end (Scheme 1). These structural features result in its low

148

absorption of light above 220 nm (Figure 1). However, the additional C=C bond



149

(C18−C19) and carbonyl group (C11−O) of SAL (and NAR) contribute to two significant

150

absorbance bands, which peak at below 230 nm and 285 nm, respectively (Figure 1). The

151

first-order rate constant of photolysis of NAR under UV irradiation was 0.47±0.01 h-1,

152

nearly identical to that of SAL (0.46 h-1, Table 2) due to their similar molecular structures.

153

The half-life of SAL in DI water with sunlight radiation (in April) was 53.3±4.1 h,

154

which is relatively long compared to other commonly used veterinary pharmaceuticals.

155

For example, the half-life of tetracycline by photolysis was around 44 min in July;28 and

156

that of sulfamethoxazole was around 10.4 h in summer.29 However, considering the fairly

157

low molar extinction coefficient of SAL (Figure 1), it suggests that SAL has a relatively

158

high quantum yield compared to other veterinary pharmaceuticals.

159

Quantum yield of SAL. The quantum yields of SAL under different light radiation are

160

shown in Table 2. The determination of photo-parameters with respect to the

161

experimental setup is described in SI Text S5 and Figure S5. The calculation of quantum

162

yield is demonstrated in SI Text S6 and Figure S6.

163

The quantum yield of SAL was similar under the three different light sources.

164

Comparing to other pharmaceuticals, the quantum yield of SAL is considerably higher.

165

For example, the quantum yields of tetracycline (0.00024−0.002),28 naproxen (0.012),30

166

diclofenac (0.031−0.22),31,32 carbamazepine (0.00013),33 levofloxacin (0.00008),34

167

amoxicillin (0.0045-0.0060),35 and sulfamethoxazole (0.02)33 are all in the range of

168

0.001−0.22 under simulated sunlight or natural sunlight, which are much lower than that

169

of SAL. On the other hand, some ketone-containing compounds, such as penton-2-one,

170

hexan-2-one, 2,2-dimethylheptan-3-one, 3,3-dimethylbutan-2-one, 2,2-dimethylhexan-3-

171

one, and 2,2,4,4-tetramethylpentan-3-one have quantum yield values of 0.27−0.71.36,37


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The similar range of quantum yields of SAL and the ketone-containing compounds

173

suggests that the photo-reactive part of SAL is likely the ketone moiety (C11−O).

174 175

Indirect photolysis of IPAs. Although MON is stable under all three light sources in

176

DI matrix, indirect photolysis of MON is expected in real water matrices via reactions

177

with reactive species generated by natural photo-sensitizers. For SAL, the observed

178

photolysis rate in real water matrix may also be influenced by different water components

179

due to various mechanisms, such as light-shielding effects, energy transfer to/from photo-

180

excited DOM, and/or reactions with reactive species generated by photo-sensitizers.

181

Nitrate and DOM are real water components particularly considered in this study,

182

because (1) they are known to have great impact on photolysis of micropollutants in

183

water systems;38-41 and (2) they are major components in WWTP effluent, agricultural

184

runoff and natural water bodies.

185

Nitrate effect. Nitrate concentrations in WWTP effluent and agricultural runoff could

186

reach 10 mg·L-1 as N or higher. In this study, 0−10 mg·L-1 nitrate as sodium nitrate was

187

spiked into DI water matrix under simulated sunlight exposure to investigate the

188

influence of nitrate on IPA photolysis. As Figure 2 shows, the presence of nitrate

189

enhanced the photodegradation of both MON and SAL. The first-order degradation rate

190

constant of MON increased linearly with increasing nitrate concentration (R2 > 0.99). The

191

presence of nitrate also enhanced the degradation of SAL under simulated sunlight

192

irradiation; however, increasing the nitrate concentration resulted in diminished

193

enhancement of the degradation rate of SAL. Because the presence of nitrate did not raise

194

the light absorbance to greater than 0.1, light-shielding effect was negligible.42 The



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195

combined results suggest that the presence of nitrate enhanced the photolysis of both

196

MON and SAL, but likely via different mechanisms.

197

Nitrate is well known to produce hydroxyl radicals under either UV or sunlight

198

exposure.43-46 Therefore, attempts were made to apply an established numerical model

199

which estimates the steady-state HO· concentration from nitrate photolysis (Equation 2)

200

in buffered DI matrix,47,48 to predict the photolysis rates of IPAs at different nitrate

201

concentrations.

202

[HO⋅]ss

∫ 2.303 ⋅ I =

λ

⋅ l ⋅ ε NO − ,λ ⋅ [NO 3− ] ⋅ Φ NO −

∑

3

3

k Si ⋅ [Si ]

(2)

203

where

204

[HO·]ss = steady-state hydroxyl radical concentration, M

205

Iλ = the irradiance received by the solution at given wavelength, Einstein·L-1·s-1

206

ελ = molar absorption coefficient of nitrate at given wavelength (M-1·cm-1)

207

l = light path length of reactor, cm

208

Φ = quantum yield of HO· production from nitrate (0.017 M·Einstein-1)

209

[Si] = concentrations of co-solutes (Si), including IPA, H2PO4-, HPO42-, HCO3- and

210

CO32-, M

211

kSi = second-order rate constants of co-solutes with HO·, M-1·s-1 (see SI Table S1)

212

After computing [HO·]ss, the indirect photolysis rate constants of IPAs in the

213

presence of simulated sunlight and nitrate were predicted by multiplying [HO·]ss with the

214

known second-order rate constants of IPAs with hydroxyl radicals.25 As Figure 2 shows,

215

the predicted rate constants agreed well with the experimental data, which suggests that

216

indirect photolysis of MON with nitrate was primarily via the reaction with HO·.


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217

However, the model underestimated the degradation rates of SAL at 0.2−1 mg·L-1

218

NaNO3 (Figure 2), which suggests reactive species other than HO· may play a part in the

219

indirect photolysis of SAL. Indeed, besides generation of HO·, the photolysis of nitrate

220

also produces several reactive nitrogen species (RNS), primarily NO2·,39,46,48 which could

221

potentially react with SAL (illustrated in SI Figure S7). Thus, the photolysis of SAL in

222

the presence of nitrate was likely partly contributed by reactions with RNS. Furthermore,

223

at an increasing nitrate concentration, the RNS concentration does not necessarily

224

increase proportionally due to the possibility of radicals combining: NO· + HO· → HNO2

225

(k = 1.0×1010 M-1·s-1), NO2· + HO· → HOONO (k = 1.3×109 M-1·s-1) and NO2·+ NO2· →

226

N2O4 (k = 4.5×108 M-1·s-1).49,50 In addition, at elevated nitrate concentrations, there would

227

be more HO· which, upon reaction with IPAs in aerated solution, would produce

228

superoxide (O2-·). The latter could inhibit photonitration reactions.51 The above

229

mechanisms could explain the non-linear dependence of the rate constant on the

230

increasing nitrate concentration (Figure 2). Thus, it is hypothesized that, at a lower nitrate

231

concentration, RNS contributed to degradation of SAL, resulting in underestimation of

232

photolysis rate by the steady-state hydroxyl radical model. In contrast, at a higher nitrate

233

concentration, RNS was self-quenched yielding a diminished increase in the photolysis

234

rate of SAL. The reaction between SAL and RNS was further supported by elucidation of

235

the transformation products (see details below).

236

DOM effect. The effect of DOM on IPA photolysis was tested under simulated

237

sunlight irradiation. SRHA (3 mg C·L-1), MPHA (3 mg C·L-1), and water extracts from

238

poultry litter (PL-extract, 30 mg C·L-1) were selected because they represent commonly

239

used model DOM (i.e., SRHA and MPHA) or DOM that IPAs will likely encounter in



240

agricultural fields (i.e., PL-extract). The UV-Vis spectra of the DOM working solutions

241

are shown in SI Figure S8. The degradation of IPAs was strongly influenced by DOM

242

(Table 2). Both MON and SAL degraded slightly faster in the presence of SRHA and

243

MPHA, compared with the rates in DI water matrix. However, PL-extract significantly

244

inhibited the photolysis of SAL.

245

The effects of DOM on the photolytic transformation of organic compounds have

246

been extensively studied and reported in the literature.52-54 In general, irradiation of DOM

247

may produce light-excited DOM (DOM*), singlet oxygen, and HO· species that can react

248

with organic pollutants in water and lead to enhanced pollutant degradation.54,55 On the

249

other hand, DOM may inhibit pollutants’ photolysis by light-shielding and

250

physical/chemical quenching effects, particularly at high DOM concentrations.56-58

251

In this study, both enhancement and inhibition of SAL photolysis by DOM were

252

observed. All three types of DOM showed absorbance greater than 0.1 at around 285 nm

253

(SI Figure S8), suggesting that light-shielding by DOM could not be neglected in the

254

experiments. The light-shielding effect was calculated to be around 14%, 21%, and 25%

255

for SRHA, MPHA and PL-extract, respectively, under the employed experimental

256

conditions. Even though light-shielding effects by the DOM were comparable, PL-extract

257

exhibited strong inhibition while SRHA and MPHA exhibited moderate enhancement on

258

SAL photolysis. This difference was likely due to different properties of the tested DOM.

259

For example, when normalizing the light absorbance of DOM from 220 to 400 nm by

260

DOM’s DOC content (A220-400·mg C-1·L-1), PL-extract had the lowest light absorbance per

261

carbon basis. This result suggests that PL-extract contained a greater portion of non-

262

photo-reactive organic compounds which might lower PL-extract’s potential to produce


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reactive species upon irradiation and compete with IPAs to react with reactive species

264

produced by photo-sensitizers. Furthermore, the inhibitory effect of PL-extract was

265

considerably greater than that attributable to light shielding, suggesting that other

266

inhibitory effects, such as physical/chemical quenching and anti-oxidation, by PL-extract

267

might also take place. Indeed, literature has shown that DOM can interact with excited

268

triplet states of ketones, leading to inhibited photolytic transformation of the ketone

269

moiety.59,60 Based on the results with PL-extract, the organic matter in runoff from PL-

270

fertilized fields is likely to exert mostly scavenging effects on IPAs’ photodegradation.

271

Photolysis in real water matrix. The photolysis of IPAs was examined in three real

272

water samples under different light irradiations, including WWTP secondary effluent

273

(under UV), simulated runoff from PL-fertilized land (under sunlight), and simulated PL

274

leachate (under sunlight). Characteristics of these water matrices are listed in Table 1.

275

Both MON and SAL photodegraded in the WWTP effluent and field runoff matrices,

276

whereas no significant loss of IPAs was detected in the PL leachate matrix for over 5

277

days. The half-lives of MON and SAL were 1.2 and 0.95 h in WWTP effluent,

278

respectively. Considering a typical UV disinfection dosage of 100 mJ·cm-2 (i.e., around 2

279

min of UV irradiation in this study), neither MON nor SAL would be efficiently removed

280

by UV treatment (less than 3% removal). Indeed, our previous study showed that, even at

281

LP UV combined with 30 mg·L-1 H2O2 AOP conditions, it required ~1000 mJ·cm-2 to

282

remove over 90% of IPAs in WWTP effluent matrix.25 In field runoff, the photolysis rate

283

constants of MON and SAL were 0.007 and 0.015 h-1 in the spring season; compared to

284

the photolysis of IPAs in DI water, the photodegradation rate of MON was significantly

285

increased, whereas the photodegradation of SAL was not enhanced.



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286

The overall photodegradation of IPAs in real water matrices is a result of both direct

287

and indirect photolysis. A steady-state model (Equation 3) was applied to predict IPA

288

photolysis rates with the assumptions that (1) nitrate was the main photo-reactive

289

constituent in the sample; and (2) DOM was taken into account for scavenging

290

HO· radicals. k obs = k direct + k indirect

291

∫

= Φ IPA/λ ⋅ (1 − 10 - A λ ) ⋅ f IPA, λ ⋅ I λ +

∫

k IPA/HO⋅ I λ ⋅ (1 − 10 - A λ ) ⋅ f NO − , λ ⋅ Φ NO − k IPA/HO⋅ ⋅ [IPA] + k DOM/HO ⋅ ⋅ [DOM] +

3

3

k HCO - /HO ⋅ ⋅ [HCO 3- ] + 3

k CO 2- /HO ⋅ ⋅ [CO 32 - ] 3

(3)

292

Notations were in the same manner as those in Equation (2). Additionally, Aλ was the

293

absorbance of water matrix at wavelength λ, and fIPA,λ and fNO3-,λ were the fractions of

294

light absorbed by IPAs and nitrate, respectively, at wavelength λ. Applying this model,

295

the observed rate constants of MON and SAL were estimated for different water matrices.

296

The predicted constants are close to the experimental results for the field runoff and PL

297

leachate matrices (Table 2). Although the model predicted the photodegradation rate of

298

MON in WWTP effluent well, it substantially overestimated the photodegradation rate of

299

SAL (Table 2). This overestimation might be caused by stronger physical/chemical

300

quenching of photo-excited ketone moiety of SAL by DOM in the WWTP effluent.

301 302

Product identification and phototransformation mechanisms. Products of SAL via

303

direct photolysis. Nine new significant peaks representing the major transformation

304

products of SAL (SAL-TPs) were observed in the direct photolysis experiments (SI

305

Figure S9), which had m/z (sodium adduct) of 265, 337, 447, 489, 491, 507, 519, 531,

306

and 805. Although all nine SAL-TPs were detected in the samples under UV, simulated


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307

sunlight and natural sunlight irradiation, the most abundant products were different for

308

each light source assuming similar MS signal response for each product (SI Figure S10).

309

Under UV irradiation, m/z 489 and 805 were the dominant SAL-TPs in the first 4 h of

310

irradiation and then were replaced by m/z 265. SAL-TPs with m/z 447, 265, 507 were

311

most abundant under simulated sunlight irradiation, whereas m/z 447 was not among the

312

major SAL-TPs in samples exposed to natural sunlight. This light-dependent generation

313

of SAL-TPs may result from two possible mechanisms. First, the absorbance of SAL

314

overlapped with UV, simulated sunlight and natural sunlight at different wavelengths (SI

315

Figure S6), which may favor specific photolytic transformation pathways to yield certain

316

dominant SAL-TPs. Indeed, dominant SAL-TPs were more similar under simulated and

317

natural sunlight, than those under LP UV. Second, the stability of SAL-TPs may be

318

different under each light source irradiation condition, which led to different dominant

319

SAL-TP species.

320

LC/MS/MS was employed to obtain more structural information of SAL-TPs. The

321

mass spectra of parent SAL and select SAL-TPs are shown in SI Figure S11. The SAL-

322

TPs with m/z of 489, 491, 507, 519 and 531 exhibited almost identical fragmentation

323

patterns below m/z 500 as that of the parent SAL. The common base peaks, m/z 431(433)

324

and 403, suggested that the products retained rings B-E, i.e., the right part of SAL

325

molecule (Scheme 1). Fragment m/z 433, instead of m/z 431, was present in some SAL-

326

TPs; this may be due to addition of an alcohol group onto C12, which rendered ring B

327

more easily to obtain a hydrogen atom from the alcohol group during fragmentation,

328

instead of eliminating a hydrogen atom from C14.61 It is also noteworthy that the



329

identified SAL-TP m/z 531 is the same as the major biotransformation product of SAL in

330

poultry litter or in soil.23,62,63

331

Fragmentation of SAL-TPs m/z 265, 337, and 447 were not successful using

332

LC/MS/MS. However, the direct photolysis of NAR also yielded products m/z 337 and

333

447, indicating that these products did not contain ring A. However, the direct photolysis

334

of NAR yielded an additional product of m/z 279, which is 14 Da increased from 265.

335

Recalling that NAR differs from SAL by a methyl group on ring A (Scheme 1), the above

336

result indicated that the SAL-TP m/z 265 was produced via cleavage on C9−C10

337

(Scheme 1), whose structure is proposed in Figure 3. SAL-TP m/z 805, along with 787

338

and 789 (observed, but not significant peaks), are named oxyl-SALs. They were likely

339

the parent SAL with one or two oxygen atoms added, forming additional ketone or

340

alcohol groups. Multiple peaks of SAL-TP m/z 805 were detected in the ion

341

chromatogram (SI Figure S11), suggesting multiple sites of the parent SAL were

342

subjected to oxygen addition. It was also found that the oxyl-SALs were photolytically

343

unstable and were degraded over time.

344

Based on the identification of major transformation products, the overall direct

345

photolysis of SAL is proposed in Figure 3. The molecular structures of SAL-TPs clearly

346

indicated that direct photolysis of SAL mainly occurred on its ketone moiety. The ketone-

347

containing compounds are subject to direct photolysis with UV or sunlight irradiation,

348

whose mechanisms were detailed in Turro et al.64 Briefly, the photo-excited ketone group

349

would form a diradical intermediate (i.e., >C=O → >C·-O·), which further transforms via

350

either α-cleavage on the ketone carbon with an adjacent carbon, or hydrogen abstraction

351

from suitable donors by the ketone oxygen atom. The latter mechanism could also lead to


Page 16 of 37

Page 17 of 37


352

formation of ROS, such as ·OOH, O2-· and ·OH, with the presence of dissolved oxygen.

353

Charge-transfer reaction via photo-excited organic compounds is another mechanism

354

which commonly produces ROS in aqueous solution. However, the electron deficient

355

oxygen on the photo-excited ketone is not likely to donate electron to the environment.

356

The generation of SAL-TPs with lower molecular weights than parent SAL was likely

357

due to the α-cleavage pathway. On the other hand, the formation of oxyl-SAL clearly

358

suggested ROS generation during direct photolysis of SAL. To further assess the self-

359

photosensitized reactions of SAL, (1) t-butanol (a ROS scavenger, 1% (v/v)) or (2) MON

360

was spiked, respectively, into DI matrix with SAL under simulated sunlight irradiation.

361

For (1), the degradation rate of SAL was decreased by 27% in the samples with t-butanol

362

and no oxyl-SAL product was observed. For (2), degradation of MON occurred in the

363

solution containing 2.5 mg·L-1 SAL, in contrast to the photostability of MON (SI Table

364

S2). These observations confirmed that ROS were generated during the direct photolysis

365

of SAL. It should be noted that the role of potential impurities from the SAL chemical

366

stock was unable to be tested in this study; thus, the possibility that some reactive species

367

might be generated by unknown impurities cannot be ruled out.

368

Photolysis products of IPAs with nitrate. Five major peaks were observed during the

369

photodegradation of MON in the presence of nitrate under simulated sunlight. They were

370

with m/z (sodium adduct) of 723, 709, 707, 691 and 549 (named MON-TPs, shown in

371

Figure 4A). The MON-TPs of m/z 723, 709, 707 and 691 were observed previously in the

372

degradation of MON under UV/H2O2 AOP conditions.25 UV at 254 nm combined with

373

H2O2 is well known to produce hydroxyl radicals as primary reactive species. Thus, the

374

similar MON degradation products observed in this study indicate that reactions with



375

hydroxyl radicals were the major mechanism for photodegradation of MON in the

376

presence of nitrate. MON-TP with m/z 549 is believed to be a secondary transformation

377

product yielded from other MON-TPs.

378

Indirect photolysis of SAL with nitrate yielded almost identical SAL-TPs with direct

379

photolysis, except that a new dominant peak with m/z (sodium adduct) 536 was observed

380

(Figure 4B). The even m/z value suggested the molecule contained an odd number of

381

nitrogen atoms. Because the parent SAL only contains C, O, and H atoms, it is likely that

382

the product was formed from reaction with RNS generated by photo-excited nitrate. To

383

test this hypothesis, indirect photolysis of SAL was conducted with the presence of 15N-

384

nitrate. The new peak with m/z 537 was observed on LC/MS, instead of m/z 536,

385

confirming that one nitrogen atom was incorporated into SAL-TP. The MS/MS fragments

386

(SI Figure S11) indicated that nitration of SAL likely occurred on C35 or C36, yielding

387

the proposed molecular structure (Figure 3). It should be pointed out that, while nitration

388

of aromatic organic pollutants via photolysis of nitrate/nitrite was observed in several

389

studies,39,65-68 the above finding is the first report on the photo-nitration of a ketone-

390

containing compound. Further efforts are required to identify the mechanism of photo-

391

nitration of a ketone-containing compound, for such a reaction may play an important

392

role in the fate of pollutants in the environment.

393 394

Toxicity tests. MON was stable in water matrix without any photosensitizers. MON-

395

TPs produced via indirect photolysis of MON, once generated, were quickly further

396

degraded by the same reactive species. In contrast, SAL-TPs produced via either direct or


Page 18 of 37

Page 19 of 37


397

indirect photolysis were relatively stable in aqueous solutions. Thus, toxicity tests were

398

performed for SAL-TPs.

399

As Figure 5 shows, the growth of the target microorganism was sensitive to the tested

400

SAL concentrations (100−400 µg·L-1). A linear regression was obtained to correlate the

401

inhibitory effect and the residual SAL concentration. The inhibitory effect of SAL

402

photolysis samples via direct photolysis or indirect photolysis with NaNO3 (under Xenon

403

lamp irradiation for 0–7 h) was almost identical to that of SAL standards, which indicated

404

that the inhibitory effect of SAL photolysis samples was contributed by the residual

405

parent SAL. Thus, SAL-TPs were not toxic to the target microorganism.

406

Indeed, the antibiotic property of IPAs is related to their ability to complex with metal

407

cations.69,70 The multiple oxygen atoms on IPAs’ cyclic ether moieties can coordinate to

408

the metal cation, and the carboxylic and alcoholic end groups can connect together to

409

assume a pseudo-cyclic conformation.69 However, the major SAL-TPs were fragments

410

from SAL generated by cleavage on the ketone moiety. As a result, SAL-TPs contained

411

fewer oxygen atoms and the molecules were more rigid, which decreased the affinity to

412

complex with metal cations in a stable pseudo-cyclic structure. Thus, SAL-TPs did not

413

present inhibitory effects in the toxicity assay. However, further work is needed to

414

systematically evaluate the impact of IPA-TPs in the environment.

415 416

ENVIRONMENTAL IMPLICATIONS

417

Among the commonly used IPAs, MON shows resistance to direct photolysis

418

whereas SAL and NAR can be degraded via direct absorption of UV or sunlight.

419

Environmental conditions can greatly affect IPAs’ photodegradation. Water near



420

agricultural fields contains significant amounts of nitrate and humic substances, which

421

will affect the photodegradation of IPAs. On the other hand, the DOM from poultry litter

422

showed strong inhibition on the direct photolysis of SAL. More than one type of IPAs

423

may co-exist in animal waste or agricultural runoff depending on feed composition.

424

Interestingly, this study showed that MON’s photolysis can be sensitized by SAL in DI

425

water via reactive species production. However, such impact from SAL may be modest in

426

the environment because SAL concentration in natural waters is expected to be low.

427

Additionally, the toxicity tests showed that photolysis of IPAs eliminated the antibiotic

428

properties against target microorganisms.

429

Because photodegradation strongly depends on light irradiance and water matrices,

430

IPA photodegradation will be most important in summer time in shallow water

431

containing low concentrations of DOM. The photolysis half-lives observed in this study

432

were 4.1 and 1.9 days for MON and SAL, respectively, in agricultural runoff matrix in

433

April. These half-lives are relatively short compared to other degradation mechanisms of

434

IPAs. For example, our previous study on the hydrolysis of IPAs showed that in mildly

435

acidic water (pH 6–7), MON was stable and the half-lives of SAL were 9–53 days.24 The

436

degradation of IPAs in the top soil fertilized with IPA-containing PL was also studied and

437

showed that SAL was stable while the half-life of MON was around 10 days.23 However,

438

considering that photodegradation primarily occurs in the top layer of water bodies, the

439

IPA photodegradation rates in the environment are expected to be slower than the

440

observed rates in this study. Nevertheless, it is evident that photodegradation of IPAs is a

441

competitive process compared to hydrolysis and biodegradation, which all contribute to

442

the environmental fate of IPAs.


Page 20 of 37

Page 21 of 37


443 444

ASSOCIATED CONTENT

445

Supporting Information. Text S1-S5, Tables S1-S2 and Figures S1−S8. This material is

446

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

447 448

ACKNOWLEDGMENTS

449

This study was supported by the U.S. Department of Agriculture Grant 2009-65102-

450

05843. The authors thank Dr. John Crittenden for providing access to simulated sunlight

451

apparatus and spectroradiometer.

452 453

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Table 1. Characteristics of real water samples DOC (mg C/L)

Nitrate (mM)

pH

HCO3(mM)

CO32(mM)

WWTP secondary effluent

0.435

1.45

7.4

1.20×10-1

1.41×10-4

Simulated field runoff from PL-fertilized land

60.55

0.71

6.5

1.51×10-2

2.24×10-6

Simulated PL leachate

2074

0.01

7.2

7.59×10-2

5.62×10-5

656 657 658

Table 2. Rate constants of MON and SAL (h-1) (R2 > 0.98) UV

Quantum yield (Φ)

Xenon

Sunlight

MON

SAL

MON

SAL

MON

SAL

ND

0.66±0.05

ND

0.66±0.03

ND

0.70±0.03

---------------------------------------------------- k Values (h-1) ----------------------------------------------ND 0.46±0.01 ND 0.24±0.01 ND 0.013±0.001 DI water a DI + Nitrateb

0.53±0.04 1.00±0.07 0.53±0.01 0.72±0.03

c

0.20±0.04 0.27±0.01

c

0.15±0.05 0.31±0.12 ND 0.08±0.02

DI + SRHA

DI + MPHA DI + PL-extractc WWTP effluentc

0.56±0.07 0.74±0.05 (1.10) (0.59) f 0.007±0.001 0.015±0.001

Field runoff d PL leachate e 659 660 661 662

ND

ND

ND

ND

a

(0.007)

(0.016)

ND (

Photodegradation of Veterinary Ionophore Antibiotics under UV and

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