Elucidating the Stimulatory and Inhibitory Effects of Dissolved Organic

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Elucidating the stimulatory and inhibitory effects of dissolved organic matter from poultry litter on photodegradation of antibiotics Kiranmayi P Mangalgiri, and Lee Blaney Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b03482 • Publication Date (Web): 27 Sep 2017 Downloaded from http://pubs.acs.org on September 27, 2017

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Environmental Science & Technology

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Elucidating the stimulatory and inhibitory effects of dissolved organic matter

2

from poultry litter on photodegradation of antibiotics

3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30

Kiranmayi P. Mangalgiri 1 and Lee Blaney 1*

1:

University of Maryland Baltimore County Department of Chemical, Biochemical and Environmental Engineering 1000 Hilltop Circle, ECS 314 Baltimore, MD 21250 USA

* Corresponding author: Lee Blaney, PhD University of Maryland Baltimore County Department of Chemical, Biochemical and Environmental Engineering 1000 Hilltop Circle, ECS 314 Baltimore, MD 21250 USA Tel: +1-410-455-8608 Fax: +1-410-455-1049 Email: [email protected]

1 ACS Paragon Plus Environment


31

ABSTRACT

32

This study examined the photolytic fate of the chlortetracycline (CTC), ciprofloxacin (CIP),

33

roxarsone (ROX), and sulfamethoxazole (SMX) antibiotics in agriculturally-relevant matrices.

34

The observed photodegradation kinetics for antibiotics in solutions containing dissolved organic

35

matter (DOM) from three poultry litter extracts were modeled to identify contributions from

36

direct and indirect photolysis. Suwannee River natural organic matter (SRN) was used as a

37

surrogate DOM standard. Poultry litter-derived DOM generated lower concentrations of reactive

38

species compared to SRN. Direct photolysis was the dominant transformation mechanism for

39

CIP, whereas CTC, ROX, and SMX were sensitized by 3DOM* and 1O2. The impacts of

40

agricultural DOM on photodegradation of antibiotics were identified in terms of pseudo-first-

41

order rate constants for formation of reactive species and second-order rate constants for reaction

42

of reactive species with DOM. Solutions containing poultry litter-derived DOM generated

43

similar levels of 3DOM* and 1O2, enhancing degradation of CTC, ROX, and SMX. The

44

reactivity of SMX was markedly different in solutions containing poultry litter DOM compared

45

to solutions with SRN, indicating that the photolytic fate of select antibiotics varies for

46

agricultural and surface water matrices. As the majority of antibiotics are consumed by animals,

47

these findings provide new insight into agriculturally-relevant transformation mechanisms and

48

kinetics.


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TOC GRAPHIC

50 51



52

1. INTRODUCTION

53

Antibiotics have been widely used in the poultry industry as prophylactics to prevent the spread

54

of infections, coccidiostats to control gut bacteria, and additives to improve feed efficiency 1.

55

Incomplete metabolism leads to excretion of antibiotics into poultry litter 2-5. For example,

56

approximately 90% of the organoarsenicals added to feed were detected in poultry litter 6.

57

Fluoroquinolone antibiotics were present in poultry litter at concentrations as high as 225

58

(mg norfloxacin) kg-1 and 1420 (mg enrofloxacin) kg-1 7. Similarly, chlortetracycline (CTC) and

59

tylosin have been measured at 20 mg kg-1 7 and 13 mg kg-1 8, respectively. Because poultry litter

60

is often applied as a soil amendment 9, 10, antibiotics are introduced to the environment through

61

animal waste.

62

Repeated land-application of antibiotic-laden poultry litter has led to changes in soil

63

microbiology 11-13, uptake in soil biota 14, and phytoaccumulation in food crops and plants 15-18.

64

Furthermore, the pseudo-persistence of antibiotics in soil has been associated with the

65

development of antibiotic resistance 19-21. In one study, 63% of poultry litter samples contained

66

Enterococcus spp. resistant to lincomycin, macrolides, and tetracyclines 2. While the number of

67

studies reporting the presence of antibiotics in the environment is increasing, the fate of

68

antibiotics in agriculturally-impacted waters has not been fully explored.

69

Solar irradiation is an important process that drives the overall fate of antibiotics in the

70

environment 22. Direct photolysis occurs when antibiotic molecules absorb light, become

71

excited, and undergo chemical transformation. Direct photolysis follows pseudo-first-order

72

kinetics 23, and the rate constant is a function of the wavelength-dependent molar absorptivity

73

and quantum yield 24. In the presence of dissolved organic matter (DOM) and other


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chromophores, direct photolysis may be inhibited due to light screening 25. For instance,

75

Wammer et al. 26 reported that 6 (mg C) L-1 of surface water DOM decreased the apparent rate

76

constant for photodegradation of fluoroquinolone antibiotics by 20-40%.

77

Antibiotics also undergo indirect photolysis through interaction with reactive species generated

78

from irradiation of DOM. Light screening (i.e., inner-filter effects) affects the generation of

79

these reactive species and results in less transformation 27. The major reactive species of concern

80

include the following: carbonate radical (CO3-•); excited triplet state DOM (3DOM*); hydrated

81

electron (e-aq); hydroxyl radical (HO•); peroxide radical (O2•-2); singlet oxygen (1O2); sulfate

82

radical (SO4-•); and, superoxide radical (O2•-) 28, 29. These mechanisms are primarily responsible

83

for photodegradation of antibiotics with low molar absorptivity or quantum yield. For example,

84

bacitracin A did not undergo significant direct photolysis after 5 h irradiation at 765 W m-2, but

85

75% transformation was achieved in the presence of 6-17 (mg C) L-1 Suwannee River natural

86

organic matter (SRN) due to reaction with 1O2 30. Similarly, Wang et al. 31 reported less than

87

10% transformation of atenolol in the absence of DOM over 50 hours of irradiation, but greater

88

than 90% degradation was observed with 20 (mg C) L-1 from Suwannee River fulvic acid

89

through 3DOM* reactions. Because the reactive species identified above also interact with

90

DOM, solutions with high DOM concentrations may inhibit photodegradation of antibiotics by

91

quenching 1O2 and 3DOM*.

92

Previous photodegradation studies have focused on surface water and wastewater effluent due to

93

concerns about antibiotic concentrations 32 and the downstream impacts on human and ecological

94

health 33. These matrices typically have low dissolved organic carbon (DOC) concentrations

95

(i.e., less than 20 mg L-1) and high transmissivity (i.e., greater than 95%). The generation and



96

quenching of reactive species at low DOM conditions, and their interaction with antibiotics, has

97

been previously reported 34. For example, degradation of sulfonamide antibiotics with 2.5 (mg

98

C) L-1 from Suwannee River and Pony Lake fulvic acids proceeds through reaction with 3DOM*

99

35

; similar mechanisms were attributed to sulfonamide transformation in real river water and

100

wastewater solutions with less than 7 (mg C) L-1. However, agricultural wastewater, lagoon

101

water, and runoff all have high DOC concentrations (i.e., greater than 50 (mg C) L-1) 36, and the

102

role of agricultural DOM at these levels has not been characterized for photodegradation of

103

antibiotics. Because DOM can also directly interact with antibiotics to form aqueous complexes

104

37-39

105

Standardized DOM sources are not available for agricultural waste. For that reason, it is difficult

106

to compare studies on the fate of antibiotics in animal waste as the variation in antibiotic

107

photoreactivity with DOM from different sources is unknown.

108

The objectives of this work were to (1) measure the generation of reactive species from poultry

109

litter-derived DOM as a function of DOC concentration (up to 140 mg L-1) and (2) deconvolute

110

the photolysis kinetics of antibiotics in simulated agriculturally-impacted water matrices to

111

determine the dominant reaction mechanisms. We focused on photodegradation of four

112

antibiotics, namely CTC, ciprofloxacin (CIP), roxarsone (ROX), and sulfamethoxazole (SMX),

113

in solutions containing DOM from three poultry litter sources or SRN. The selected antibiotics

114

stem from four classes (i.e., tetracyclines, fluoroquinolones, organoarsenicals, and sulfonamides,

115

respectively) used in the global poultry industry 40-42 and previously detected in poultry litter 7, 43-

116

48

117

these chemicals are still widely used around the world 51. CIP, CTC, and SMX have been

118

identified as “critically important to human health” by the World Health Organization 42. An

, the photolytic fate of antibiotics in agriculturally-relevant conditions is complex.

. ROX and other organoarsenicals have been banned in Europe and North America 49, 50, but


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improved understanding of the photolytic fate of antibiotics in agriculture-impacted waters is

120

necessary to mitigate environmental and human health concerns associated with antibiotic use in

121

intensive agriculture.

122

2. METHODS AND MATERIALS

123

2.1 Chemicals and Reagents

124

CIP (> 99%), CTC (> 80%), ROX (> 95%), and SMX (> 95%) were purchased from Sigma-

125

Aldrich (St. Louis, MO). The following scavengers and probe compounds were secured from

126

Fisher Scientific (Pittsburgh, PA) for competition kinetics analysis: 2,4,6-trimethylphenol (TMP,

127

probe for 3DOM*), furfuryl alcohol (FFA, probe for 1O2), para-chlorobenzoic acid (pCBA, probe

128

for HO•), sodium azide (NaN3, scavenger for HO• and 1O2), sorbic acid (scavenger for 3DOM*),

129

and tert-butanol (t-BuOH, scavenger for HO•). Buffers were constructed using formic acid and

130

monobasic, dibasic, and tribasic sodium phosphate salts from Fisher Scientific. The Rose Bengal

131

1

132

Humic Substances Society (Denver, CO). All solutions were generated in deionized (DI) water

133

(18 MΩ) produced from an in-house system that employs sequential adsorption, ion exchange,

134

reverse osmosis, and ultraviolet (UV) disinfection processes (Neu-Ion Systems; Baltimore, MD).

135

2.2 DOM Characterization

136

Poultry litter samples from three commercial farms in the Chesapeake Bay watershed were used

137

as DOM sources. Litter samples were oven-dried at 40 °C, sieved (1.19 mm), and homogenized.

138

Poultry litter extracts (PLEs) were prepared at room temperature by adding 40 g of the dried,

139

homogenized poultry litter to 1 L of DI, shaking for 20 minutes at 250 rpm, and centrifuging at

140

14,000 g for 45 minutes at 26 °C. Aliquots (1 mL) of 0.45 µm filtered extracts were stored

O2 sensitizer was obtained from Fisher Scientific. SRN was acquired from the International



141

at -20 °C, and working solutions were diluted as needed. These extracts were constructed to

142

simulate DOM leaching from poultry litter during irrigation and/or rainfall events. A standard

143

SRN solution was made by adding 2.5 g of SRN to 1 L of DI. The resulting PLE and SRN stock

144

solutions exhibited similar absorbance in the 310 – 410 nm range.

145

Total organic carbon (TOC) concentrations were measured using a Shimadzu TOC-L instrument

146

(Columbia, MD). Samples were typically diluted 100 times before measuring UV absorbance

147

with (i) a 1-cm quartz cuvette in the Evolution 600 spectrophotometer (Thermo; Waltham, MA)

148

or (ii) UV-transparent 96-well plates in the Eon microplate reader (BioTek; Winooski, VT).

149

Fluorescence excitation-emission matrices (EEMs) were recorded for 100× diluted samples in 1-

150

cm quartz cuvettes using a Cary Eclipse fluorescence spectrophotometer (Varian; Walnut Creek,

151

CA).

152

2.3 Chemical Analysis

153

The concentrations of the four antibiotics and the three reactive species probe compounds were

154

measured by liquid chromatography with diode array detection and tandem mass spectrometry

155

(LC-DAD-MS/MS; Thermo UltiMate 3000 with Quantum Access Max). LC-MS grade water

156

and methanol (both with 0.1% formic acid) were used to generate the mobile phase. The

157

analytes were separated on a Waters Symmetry C18 column (2.1×150 mm, 3.5 µm). Details on

158

the elution gradient, analyte retention times, DAD settings, and MS/MS parameters are provided

159

in Figure S1, Table S1, and Text S1 of the Supporting Information (SI).

160


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2.4 Photodegradation Experiments

162

Photodegradation experiments were conducted in a merry-go-round Rayonet RMR 600 reactor

163

(Southern New England Ultraviolet Inc.; Branford, CT) equipped with bulbs emitting in the

164

310 – 410 nm range (see Figure S2 in the SI). The average incident photon flux of the system

165

was calculated to be 2.51×10-5 Ein L-1 s-1 using the ferrioxalate actinometer 52, 53 with an average

166

quantum yield of 1.23 mol Ein-1 54. Experimental solutions containing 1 mg L-1 antibiotic,

167

variable DOC (from SRN or PLE stock solutions), and 10 mM phosphate buffer (pH 6.8, based

168

on the natural pH of the PLEs) were added to 10-mL quartz tubes and irradiated. Aliquots of

169

100 µL were extracted at pre-determined times and diluted to 1 mL with 0.1% formic acid prior

170

to analysis by LC-DAD-MS/MS. To determine steady state concentrations of reactive species,

171

TMP, FFA, and pCBA were added at 25, 25, and 1 mg L-1, respectively. To selectively scavenge

172

reactive species, NaN3, sorbic acid, and t-BuOH were added at concentrations of 100, 50, and 10

173

mg L-1, respectively. When determining the second-order reaction rate constants for antibiotics

174

with 1O2, Rose Bengal was dosed at 25 mg L-1.

175

The observed photodegradation kinetics for all antibiotics were fit to a time-based,

176

pseudo-first-order model (Eq. 1).

ln

[AB]

= [AB]

177

(Eq. 1)

178

The observed time-based rate constants ( ) for photodegradation of antibiotics were

179

calculated as the slope of the natural logarithm of the normalized antibiotic concentration



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180

([AB]t / [AB]o) plotted against time, t. The overall was comprised of contributions from

181

direct photolysis and reactions with HO•, 1O2, and 3DOM*, as described in Eq. 2.

[HO •] +

= , + •,

,

O

+

$!"∗ , [ 'DOM ∗ ]

= , + ,• + , + , $ !"∗

182

(Eq. 2)

183

In Eq. 2, S is the screening correction, , is the pseudo-first-order rate constant for direct

184

is the second-order rate constant for antibiotic reaction with photolysis of the antibiotic, (,

185

reactive species i, and ,( is the observed rate constant for direct photolysis or reaction with

186

HO•, 1O2, or 3DOM*; the ss subscript indicates steady state concentration.

187

In the absence of DOM, was interpreted as , . The value for S was dependent on the

188

DOM matrix (see Text S2 in the SI) and varied from 0.99 to 0.59 for solutions containing 2-140

189

(mg C) L-1 for the PLEs and 0.5-40 (mg C) L-1 for SRN. When available, second-order rate

190

constants for reaction of antibiotics with reactive species were collected from literature;

191

otherwise, these rate constants were experimentally measured. Steady state concentrations of

192

reactive species were obtained using the aforementioned probe molecules. Expressions for the

193

mechanism-specific rate constants are provided in Eq. 3-6 as a function of DOM concentration.

194

Detailed derivations of these expressions are available in Text S3 (see SI).

,

=

), $∗ $∗ ,!"

1 + + , $∗

[DOM]


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195

(Eq. 3)

,• =

[DOM]

),•

[AB] + + •,!" •, [DOM] +

,•

•,

196

(Eq. 4)

,

), [DOM]

=

[AB] + +

,!"

,

[DOM] +

,

,

197

(Eq. 5)

= , $ !"∗

), $!"∗ [DOM] $!"∗ ,!"

[AB] + + $

!"∗ ,

[DOM] +

,

$!"∗

$!"∗ ,

198

(Eq. 6)

199

In Eq. 3-6, ),( is the rate constant for the reaction governing formation of excited antibiotics

200

(3AB*) and reactive species (i.e., HO•, 1O2, and 3DOM*), + , $∗

201

3

$

∗ ,!"

is the preference ratio for

AB* quenching to ground state AB by DOM over direct photolysis of 3AB* (defined as

202

$∗ ,!" / , ), + (,!" $ (, is the preference ratio for reaction of reactive species with DOM ∗

203

over their reaction with an antibiotic (defined as (,!" / (, ), and ,( is the rate constant

204

for deactivation of a reactive species by water. DOM was quantified as DOC concentration. The

205

overall photodegradation kinetics of individual antibiotics in solutions with DOM were fit to the



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206

expanded form of Eq. 2 (substitution of Eqs. 3-6) using a least squares approach in OriginPro

207

(OriginLab Corp.; Northampton, MA) and Microsoft Excel. Multiple starting points were

208

employed for each antibiotic to ensure that the global minimum was reported and that the

209

corresponding rate constants represented the best fit to experimental data.

210

3. RESULTS AND DISCUSSION

211

3.1 Generation of Reactive Species from Poultry Litter

212

Physicochemical properties of the DOM sources are reported in Table 1. The absorbance spectra

213

for 310 – 410 nm were similar for the four DOM stock solutions (see Figure S2a in the SI),

214

resulting in comparable screening corrections (see Table S2 in the SI). Fluorescence analysis

215

(Figure S2c-e in the SI) of the PLEs revealed terrestrial humic-, microbial humic-, and

216

protein-like signatures 55, 56. The fluorescence intensity of the protein-like pool in the PLEs was

217

significantly greater than that of SRN (Figure S2b in the SI), which mostly consists of fulvic

218

acid- and humic-like components 57. The composition of SRN agrees with the higher SUVA254

219

and lower E2/E3 ratios reported in Table 1 and indicates a fairly aromatic DOM with a higher

220

molecular weight distribution compared to the poultry litter-derived DOM 58. The observed

221

compositional differences suggest that DOM from poultry litter and SRN will exhibit variable

222

generation of, and reaction with, reactive species during photolysis 59-61.

223

Insert Table 1

224

The 1O2 concentrations were measured using the FFA probe compound (

225

M-1s-1 62). For all DOM sources, steady state concentrations of 1O2 (see Figure 1a) were in the

226

10-13 – 10-12 M range, in agreement with previous reports for surface waters 29. For the PLEs,


,--

= 1.0×108

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227

1

228

compositions 56. Steady state concentrations of 1O2 in the SRN solutions were higher than those

229

in the PLE-containing solutions. For example, 37 (mg C) L-1 resulted in 1.19×10-12 M 1O2 for the

230

SRN matrix, nearly double that of PLE-2 (i.e., 6.18×10-13 M 1O2). As predicted by previous

231

studies 63, the steady state concentration of 1O2 increased linearly for all sources at low DOC

232

levels (i.e., less than 30 mg L-1). Above this threshold, 1O2 reaction with DOM caused a

233

noticeable change in the 1O2 yield.

O2 concentrations were comparable for specific DOC levels, potentially due to the similar PLE

234

Insert Figure 1

235

When these data were plotted against the screening factor, the steady state concentrations of 1O2

236

were similar for all four DOM sources (Figure 1c). For a screening factor of 0.95-0.96, the 1O2

237

concentrations were 1.96×10-13 M and 2.14×10-13 M for SRN and PLE-1, respectively. This

238

observation suggests that the quantum yield for 1O2 formation was similar for SRN and the

239

PLEs, and the differences in steady state 1O2 concentrations at higher DOM levels are associated

240

with the specific rate of light absorption of the DOM sources. Given the lower SUVA254 and

241

higher E2/E3 of the PLEs, higher 1O2 concentrations were expected based on previous research

242

reporting higher 1O2 quantum yields for DOM with lower molecular weight distributions 64. In

243

this case, the inherent differences in the DOM composition (e.g., the high fulvic- and humic-like

244

character of SRN compared to the relatively high protein-like fluorescence of the PLEs) or co-

245

extracted inorganic ions may cause different trends in 1O2 quantum yield or scavenge some of

246

the produced 1O2.

247

The observed steady state concentrations of 3DOM* were determined using the TMP probe

248

molecule (

$!"∗ ,."/ = 3.0×109 M-1s-1 65), while accounting for TMP reaction with 1O2 13 ACS Paragon Plus Environment


249

(

250

associated with 3DOM* reaction, which is the expected mechanism for DOM reactivity with

251

antibiotics. Like 1O2, 3DOM* concentrations increased in a linear fashion at low DOC

252

concentrations; however, 3DOM* levels plateaued around 30 (mg C) L-1 for PLE-0 and PLE-2,

253

presumably due to reaction with DOM. Wenk et al. 67 also reported scavenging of triplet state

254

sensitizers by aquatic and terrestrial DOM at 20-70 (mg C) L-1. The three PLEs generated

255

different 3DOM* concentrations for the same DOC levels (Figure 1b) and screening factors

256

(Figure 1d). For high DOM (i.e., > 30 (mg C) L-1), 3DOM* concentrations in the PLE-0 solution

257

were approximately 50% of those for PLE-2, suggesting that formation of 3DOM* was slower for

258

PLE-0 or the reaction of 3DOM* with DOM was faster for PLE-0. Steady state concentrations of

259

3

260

attributed to the greater presence of fulvic-like molecules in SRN, as these compounds have been

261

associated with greater steady state concentrations of 3DOM* 68. A parallel factor analysis of

262

EEMs for two of the PLEs used in this study did not identify fulvic acid components 56.

263

No degradation of pCBA was observed in these experiments, suggesting that HO• was not

264

generated at an appreciable concentration (i.e., [HO•]ss < 10-18 M for our experimental

265

conditions). Therefore, HO• reaction mechanisms in Eq. 2 were not considered during fitting of

266

observed rate constants.

267

3.2 Direct Photolysis of Antibiotics

268

for the four antibiotics ranged over three orders of magnitude: CIP, At pH 6.8, ,

269

7.47 (± 0.40)×10-3 s-1; CTC, 2.84 (± 0.49)×10-4 s-1; SMX, 6.62 (± 0.46)×10-5 s-1; and, ROX,

270

1.33 (± 0.27)×10-6 s-1. The fluence-based, pseudo-first-order rate constant and average molar

,."/

= 6.7×107 M-1s-1 66). The TMP probe evaluates the electron transfer pathway

DOM* for the SRN matrix were higher than those for the PLEs. These findings may be


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271

absorptivity at 310 – 410 nm for each antibiotic are provided in Table 2. More details on these

272

calculations can be found in Text S4 (see SI). The photoreactivity of the antibiotics did not

273

directly correlate to their molar absorptivity. For example, CIP and ROX have similar

274

absorbance in the 310 – 410 nm range, but CIP is three orders of magnitude more reactive than

275

ROX. These findings indicate that the quantum yields vary significantly between antibiotic

276

classes. Self-sensitization of antibiotics during direct photolysis was not found to be significant

277

(see Figure S3 in the SI). Specific reaction mechanisms for direct photolysis have been reported

278

elsewhere for CIP 69, 70, CTC 71, and SMX 72, 73. The indirect photolysis of antibiotics due to

279

irradiation of DOM from PLE-0, PLE-1, PLE-2, and SRN at agriculturally-relevant

280

concentrations is discussed below.

281

Insert Table 2

282

3.3 Effect of DOM on Antibiotic Photodegradation

283

The net differences in the observed pseudo-first-order rate constants for CIP, CTC, ROX, and

284

SMX photodegradation in the presence/absence of DOM are shown in Figure 2. In Figure 2,

285

positive and negative differences imply sensitization and inhibition, respectively, of antibiotic

286

degradation by DOM. The observed rate constants for the four antibiotics showed different

287

trends with DOC concentration. For instance, the observed rate constant for CIP consistently

288

decreased with increasing DOC for all sources; however, ROX degradation was enhanced over

289

the same DOC gradient. The observed rate constant for CTC degradation exhibited marginal

290

changes for the examined DOC levels, although differences were apparent between DOM

291

sources. Photodegradation of SMX was dependent on both DOM source and concentration. For

292

example, the observed rate constant for SMX in the PLE-0 matrix increased steadily from 1000×



293

to 100× dilution, before decreasing to the approximate magnitude for direct photolysis. For the

294

SRN solution, observed rate constants for SMX degradation decreased steadily with DOC.

295

These trends can be attributed to the dominant degradation mechanisms, which depend on

296

antibiotic structure, and DOM generation of, and reaction with, reactive species.

297

Insert Figure 2

298

Dominant photodegradation mechanisms for the four antibiotics of interest were established by

299

selective reactive species scavenging experiments performed in solutions containing 9.2 (mg C)

300

L-1 SRN and 1 mg L-1 antibiotic. CIP photodegradation was not significantly enhanced by DOM,

301

but the other antibiotics were sensitized as follows: ROX, 1O2; SMX, 3DOM*; and, CTC, 3DOM*

302

and 1O2 (see Figure S4 in the SI). The following subsections expand on the photoreactivity of

303

individual antibiotics in the various DOM matrices.

304

3.3.1 Ciprofloxacin – Screening and Quenching of 3CIP* to CIP by DOM

305

DOM, regardless of source, inhibited CIP transformation. Previous studies have reported

306

suppression of fluoroquinolone photolysis in surface water and wastewater effluent 26, 69 due to

307

screening effects. Given the low reactivity of CIP with 1O2 and 3DOM* 74, direct photolysis was

308

the dominant photodegradation mechanism. The time-based degradation trends for CIP in

309

solutions containing 3 – 123 (mg C) L-1 from PLE-0 are provided in Figure S5 (see SI). Figure

310

3a shows the observed rate constant for CIP plotted against DOC concentration for all DOM

311

sources. The low for CIP at higher DOC concentrations was not accounted for by screening

312

corrections alone (see Figure S6 in SI). These results suggest that DOM quenched 3CIP* to

313

ground state CIP, as indicated by Reaction S.4 in Text S3 of the SI. Similar quenching


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314

mechanisms were reported by Wenk et al. 67 for solutions containing 22 – 72 (mg C) L-1 from

315

various DOM sources.

316

Insert Figure 3 $

01/∗ ,!"

317

The observed photolysis kinetics of CIP were fit to Eq. 3. The preference ratio +

318

reported in Table 3 for the four DOM sources. These values were comparable for all three PLEs,

319

indicating that the reaction of poultry litter-derived DOM with 3CIP* is fairly consistent. This

320

result likely stems from the similar composition of the three PLEs. The more aromatic SRN

321

matrix demonstrated faster quenching of 3CIP* to CIP. Porras et al. 75 reported that CIP

322

photolysis was inhibited in the presence of natural organic matter from a Nordic reservoir. This

323

report, in combination with our findings, suggests that DOM composition plays a critical role in

324

the fate of the 3CIP* species.

325

, $01/∗

is

Insert Table 3

326

Even though CIP photoreactivity was suppressed by DOM through screening effects and

327

quenching of 3CIP* to CIP, the in the current study remained in the 10-4 s-1 range, even at

328

the highest DOC concentrations. This observed rate constant was 1-2 orders of magnitude

329

higher than those calculated for other antibiotics in this study. These findings demonstrate that

330

CIP is effectively photodegraded in natural sunlight, even in agriculturally-impacted waters with

331

high DOC.



Page 18 of 42

332

3.3.2 Roxarsone – Reaction with 1O2

333

Selective reactive species scavenging experiments indicated that ROX reaction with 1O2 was the

334

dominant photodegradation mechanism (see Text S5 and Figure S7 in the SI). Eq. 3 and Eq. 5

335

were, therefore, used to explain the observed photodegradation trends. The second-order rate

336

constant for reaction of ROX with 1O2 was calculated to be 3.05 (±0.36)×107 M-1 s-1 (see Text S6

337

in the SI). The magnitude of this rate constant, along with the measured steady state

338

concentrations, confirmed the importance of the 1O2 mechanism for ROX degradation.

339

Figure 3b shows the observed rate constant for ROX photodegradation in solutions with different

340

DOM sources. In general, the ROX transformation kinetics increased with DOC concentration

341

for all sources, although some inhibitory effects were observed at low concentrations of PLE-0

342

and PLE-2 due to screening and DOM quenching of 3ROX* to ROX. Quenching of 3ROX* to

343

ROX varied with DOM source, as indicated by the calculated values for + , $23∗

344

2.2×106 Mc-1 for SRN; 2.2×103 Mc-1 for PLE-0; 2.2×102 Mc-1 for PLE-1; and, 6.5×104 Mc-1 for

345

PLE-2. These findings verified the different behaviors of the DOM sources at low DOC levels.

346

At high DOM concentrations, the overall contribution of the 1O2 reaction increased (see Figure

347

S8 in SI). As expected from Figure 1a, the rate constant for formation of 1O2 was higher for

348

SRN ( ), = 2.5×10-4 s-1) compared to the PLEs ( ), = 2.7 – 11×10-6 s-1). The rate constants

349

for reaction of bulk DOM with 1O2 were source-dependent. For instance, the + ,23 ratio for

350

ROX varied in the range of 0.3 – 59 M Mc-1, indicating that DOM differentially competed with

351

ROX for reaction with 1O2.

$

23∗ ,!"

:

,!"


Page 19 of 42


352

3.3.3 Chlortetracycline – Reaction with 1O2 and 3DOM*

353

Self-sensitized photodegradation of tetracyclines by 1O2 has been reported 76, but this mechanism

354

was not found to be significant for the experimental conditions examined here (see Text S3 and

355

Figure S3 in the SI). Reactive species scavenging studies (see Text S5 and Figure S7 in the SI)

356

indicated that CTC reactions with 3DOM* and 1O2 were important in the presence of 9.2 (mg C)

357

L-1 of SRN. Previous authors have described the role of 1O2 in CTC phototransformation 77, 78.

358

Figure 3c shows the variation of for CTC photodegradation as a function of DOM source

359

and concentration. The observed rate constant was fit using Eq. 3, Eq. 5, and Eq. 6.

360

The overall contribution of 1O2 to CTC degradation was low (see Figure S9 in the SI) due to

361

slow reaction kinetics (see Table 3). The net inhibition of CTC degradation in solutions

362

containing PLE-1 at DOC concentrations less than 20 mg L-1 was mostly associated with light

363

screening and quenching of 3CTC* to CTC by DOM. At higher DOM concentrations, reaction

364

with 3DOM* dominated the overall degradation of CTC. The preference ratio, + $!"∗ ,0.0 , was

365

higher for SRN than the PLEs, indicating that 3DOM* originating from PLE and SRN showed

366

source-dependent reactivity with DOM and CTC. The second-order rate constants obtained for

367

reaction of CTC with 3DOM* from each source were as follows: 1.5×109 Mc-1s-1 for PLE-1;

368

5.7×109 Mc-1s-1 for PLE-2; 7.9×109 Mc-1s-1 for SRN; and, 1.2×1010 Mc-1s-1 for PLE-0.

369

3.3.4 Sulfamethoxazole – Reaction with 3DOM*

370

Even though SMX does not absorb much light above 325 nm, the contribution of direct

371

photolysis is significant due to the high quantum yield (i.e., 0.50 ± 0.09 at pH 5.3 for solar

372

irradiation 79). Boreen et al. 79 reported that SMX does not react with 1O2. Therefore, any

373

enhancement in the photodegradation rate of SMX in DOM-containing solutions is expected to

$


!"∗ ,!"


374

occur through 3DOM* mechanisms. Similar conclusions have been reported for

375

photodegradation of SMX in wastewater 35, 80. Figure 3d shows the effect of DOM on the

376

observed rate constant for SMX, which was fit using Eq. 3 and Eq. 6.

377

DOM content and type had complex effects on the observed SMX rate constant. For solutions

378

containing SRN, consistently decreased with increasing DOC due to the low

$!"∗ ,4"3

379

(7.2×107 M-1 s-1). At low DOM levels, 3SMX* was quenched to ground state SMX by DOM.

380

Although SMX showed low reactivity with 3DOM* from SRN, reaction with 3DOM* was

381

dominant over direct photolysis at high DOM concentrations for all sources (see Figure S10 in

382

the SI). This finding supports sensitization of SMX at high DOC concentrations. The

383

$!"∗ ,4"3 values differed for each PLE source and were determined to be 2.9×109, 1.1×108,

384

and 1.0×109 M-1 s-1 for PLE-0, PLE-1 and PLE-2, respectively.

385

The enhanced photoreactivity of SMX in the PLEs may be associated with the microbial humic-

386

like components. Previous studies have indicated that 3DOM* from autochthonous sources,

387

including wastewater effluent and lake water, enhanced the photodegradation of sulfa-drugs 35, 80

388

compared to 3DOM* from Suwannee River. This effect may be associated with greater

389

antioxidant properties of SRN or greater oxidation potential of non-terrestrial sources 65.

390

3.4 Environmental Significance

391

In this study, direct and indirect photolysis kinetics and mechanisms have been analyzed for four

392

antibiotics at agriculturally-relevant conditions using organic matter extracted from three poultry

393

litters. While previous studies have described indirect photolysis of contaminants of emerging

394

concern in surface waters with low DOC, this study shows that antibiotic photodegradation also


Page 20 of 42

Page 21 of 42


395

occurs at the high DOM levels characteristic of agriculturally-impacted waters. The calculated

396

parameters associated with formation and quenching of reactive species in the PLE and SRN

397

solutions indicate that agriculturally-derived DOM does not impact photodegradation in the same

398

way as common surrogate materials like SRN. These results suggest that the photochemistry of

399

antibiotics in surface water significantly differs from agricultural waters. As more than 70% of

400

antibiotics sold in the United States are used in agriculture 81, the fundamental photodegradation

401

mechanisms and kinetics reported in this study constitute important advances in understanding

402

antibiotic fate in the environment. To gain further insight into the photodegradation of

403

antibiotics in agricultural systems, future studies should investigate real agriculturally-impacted

404

waters with mixed DOM sources.

405

The observed photoreactivity varied for each antibiotic; furthermore, the impacts of DOM source

406

and concentration manifested in different ways for each antibiotic. Degradation of CIP was

407

inhibited due to screening effects and quenching of the 3CIP* intermediate to ground state CIP by

408

DOM. CTC, ROX, and SMX showed varying degrees of sensitization due to selective reactivity

409

with 3DOM* and 1O2. Previous studies have reported phototransformation of antibiotics in

410

engineered systems 82-85 causes formation of antimicrobially-active transformation products. The

411

effects of agricultural DOM on these reactions, namely the differences in reaction pathway for

412

direct photolysis and 1O2 and 3DOM* mediated processes, is a critical area for future study.

413

Given the global use of diverse antibiotics to raise food-producing animals, the results of this

414

study highlight the need to determine the photolytic fate of antibiotics in waste management

415

systems, such as anaerobic lagoons, and agricultural runoff. Findings from this study also have

416

implications for the impact of run-off from animal-derived waste streams on surface water,



417

wherein mixed DOM sources may differentially affect the fate of antibiotics and other

418

contaminants of emerging concern, including hormones, pesticides, and herbicides.

419

SUPPORTING INFORMATION

420

Method details; analysis and fitting of kinetic model; determination of second-order rate constant

421

for roxarsone with singlet oxygen; supporting tables and figures

422

ACKNOWLEDGEMENTS

423

We gratefully acknowledge funding from NSF CHE 1508090 and CBET 1510420.


Page 22 of 42

Page 23 of 42


424

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677



678

List of Tables

679

Table 1.

Properties of 100× diluted PLE and SRN stock solutions.

680

Table 2.

Physicochemical properties of the four antibiotics of concern.

681

Table 3.

Rate constants and preference ratios for direct and indirect photolysis of the four

682

antibiotics of concern in the four DOM matrices.

683 684

List of Figures

685

Figure 1.

Steady state concentrations of (a, c) 1O2 and (b, d) 3DOM* produced during

686

irradiation of solutions containing DOC from SRN, PLE-0, PLE-1, and PLE-2

687

plotted as a function of (a, b) DOC concentration and (c, d) screening factor.

688

Figure 2.

689 690

Effect of DOM source and concentration on the observed transformation kinetics of (a) CIP, (b) ROX, (c) CTC, and (d) SMX.

Figure 3.

Effects of SRN, PLE-0, PLE-1, and PLE-2 (columns i, ii, iii, and iv, respectively) on

691

the photodegradation of CIP, ROX, CTC, and SMX (rows a, b, c, and d,

692

respectively)

693


Page 36 of 42

Page 37 of 42


694

Table 1.

695 696

Screening DOC UV254 SUVA254 E2/E3 -1 -1 factor (mg L ) (cm ) (L (mg C)-1 m-1) a 0.794 ± 0.008 4.72 ± 0.20 9.20 ± 1.73 0.491 ± 0.005 5.84 ± 1.18 SRN 0.843 ± 0.008 5.36 ± 0.99 30.97 ± 2.34 0.390 ± 0.004 1.32 ± 0.12 PLE-0 0.847 ± 0.009 5.74 ± 1.25 28.32 ± 3.82 0.384 ± 0.004 1.48 ± 0.28 PLE-1 0.851 ± 0.009 5.39 ± 0.93 35.02 ± 6.42 0.349 ± 0.004 1.09 ± 0.24 PLE-2 a: The error values for all entries represent 95% confidence intervals obtained from triplicate measurements

Properties of 100× diluted PLE and SRN stock solutions.

DOM



697

Table 2. Antibiotic

Physicochemical properties of the four antibiotics of concern. Molecular weight (g mol-1)

a

Chemical structure

+

NH2 N

CIP

N

331.35 O F O

-

NH

OH

-

478.88 O OH

O

O

O

-

As O

-

O N O H2N

-

O

N

O

253.28

1.02 (± 0.05)×10-3

3.33 86 7.55 9.33

5.91 (± 0.14)×103

3.85 (± 0.65)×10-5

3.45 83 5.95 9.15

2.12 (± 0.07)×103

1.80 (± 0.37)×10-7

1.85 86 5.65

7.59 (± 0.11)×10-2

9.01 (± 0.63)×10-6

NH 2

HO

+

SMX

2.47 (± 0.15)×103

OH

-

O

263.03

3.01 86 6.14 8.70 10.58

+

O

ROX

;

Elucidating the Stimulatory and Inhibitory Effects of Dissolved Organic

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