Elucidating the Stimulatory and Inhibitory Effects of ... - ACS Publications

Article Cite This: Environ. Sci. Technol. 2017, 51, 12310-12320

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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*,† †

Department of Chemical, Biochemical and Environmental Engineering, University of Maryland Baltimore County, 1000 Hilltop Circle, ECS 314, Baltimore, Maryland 21250, United States S Supporting Information *

ABSTRACT: This study examined the photolytic fate of the chlortetracycline (CTC), ciprofloxacin (CIP), roxarsone (ROX), and sulfamethoxazole (SMX) antibiotics in agriculturally relevant matrices. The observed photodegradation kinetics for antibiotics in solutions containing dissolved organic matter (DOM) from three poultry litter extracts was modeled to identify contributions from direct and indirect photolysis. Suwannee River natural organic matter (SRN) was used as a surrogate DOM standard. Poultry litter-derived DOM generated lower concentrations of reactive species compared to SRN. Direct photolysis was the dominant transformation mechanism for CIP, whereas CTC, ROX, and SMX were sensitized by 3DOM* and 1O2. The impacts of agricultural DOM on photodegradation of antibiotics were identified in terms of pseudo-first-order rate constants for formation of reactive species and second-order rate constants for reaction of reactive species with DOM. Solutions containing poultry litter-derived DOM generated similar levels of 3DOM* and 1O2, enhancing degradation of CTC, ROX, and SMX. The reactivity of SMX was markedly different in solutions containing poultry litter DOM compared to solutions with SRN, indicating that the photolytic fate of select antibiotics varies for agricultural and surface water matrices. As the majority of antibiotics are consumed by animals, these findings provide new insight into agriculturally relevant transformation mechanisms and kinetics.

1. INTRODUCTION Antibiotics have been widely used in the poultry industry as prophylactics to prevent the spread of infections, coccidiostats to control gut bacteria, and additives to improve feed efficiency.1 Incomplete metabolism leads to excretion of antibiotics into poultry litter.2−5 For example, approximately 90% of the organoarsenicals added to feed were detected in poultry litter.6 Fluoroquinolone antibiotics were present in poultry litter at concentrations as high as 225 (mg norfloxacin) kg−1 and 1420 (mg enrofloxacin) kg−1.7 Similarly, chlortetracycline (CTC) and tylosin have been measured at 20 mg kg−17 and 13 mg kg−1,8 respectively. Because poultry litter is often applied as a soil amendment,9,10 antibiotics are introduced to the environment through animal waste. Repeated land-application of antibiotic-laden poultry litter has led to changes in soil microbiology,11−13 uptake in soil biota,14 and phytoaccumulation in food crops and plants. 15−18 Furthermore, the pseudopersistence of antibiotics in soil has been associated with the development of antibiotic resistance.19−21 In one study, 63% of poultry litter samples contained Enterococcus spp. resistant to lincomycin, macrolides, and tetracyclines.2 While the number of studies reporting the presence of antibiotics in the environment is increasing, the © 2017 American Chemical Society

fate of antibiotics in agriculturally impacted waters has not been fully explored. Solar irradiation is an important process that drives the overall fate of antibiotics in the environment.22 Direct photolysis occurs when antibiotic molecules absorb light, become excited, and undergo chemical transformation. Direct photolysis follows pseudo-first-order kinetics,23 and the rate constant is a function of the wavelength-dependent molar absorptivity and quantum yield.24 In the presence of dissolved organic matter (DOM) and other chromophores, direct photolysis may be inhibited due to light screening.25 For instance, Wammer et al.26 reported that 6 (mg C) L−1 of surface water DOM decreased the apparent rate constant for photodegradation of fluoroquinolone antibiotics by 20−40%. Antibiotics also undergo indirect photolysis through interaction with reactive species generated from irradiation of DOM. Light screening (i.e., inner-filter effects) affects the generation of these reactive species and results in less transformation.27 The major reactive species of concern include the following: Received: Revised: Accepted: Published: 12310

July 9, 2017 September 24, 2017 September 27, 2017 September 27, 2017 DOI: 10.1021/acs.est.7b03482 Environ. Sci. Technol. 2017, 51, 12310−12320

Article

Environmental Science & Technology carbonate radical (CO3•−); excited triplet state DOM (3DOM*); hydrated electron (e−aq); hydroxyl radical (HO•); peroxide radical (O2•2−); singlet oxygen (1O2); sulfate radical (SO4•−); and, superoxide radical (O2•−).28,29 These mechanisms are primarily responsible for photodegradation of antibiotics with low molar absorptivity or quantum yield. For example, bacitracin A did not undergo significant direct photolysis after 5 h irradiation at 765 W m−2, but 75% transformation was achieved in the presence of 6−17 (mg C) L−1 Suwannee River natural organic matter (SRN) due to reaction with 1O2.30 Similarly, Wang et al.31 reported less than 10% transformation of atenolol in the absence of DOM over 50 h of irradiation, but greater than 90% degradation was observed with 20 (mg C) L−1 from Suwannee River fulvic acid through 3DOM* reactions. Because the reactive species identified above also interact with DOM, solutions with high DOM concentrations may inhibit photodegradation of antibiotics by quenching 1O2 and 3DOM*. Previous photodegradation studies have focused on surface water and wastewater effluent due to concerns about antibiotic concentrations32 and the downstream impacts on human and ecological health.33 These matrices typically have low dissolved organic carbon (DOC) concentrations (i.e., less than 20 mg L−1) and high transmissivity (i.e., greater than 95%). The generation and quenching of reactive species at low DOM conditions, and their interaction with antibiotics, have been previously reported.34 For example, degradation of sulfonamide antibiotics with 2.5 (mg C) L−1 from Suwannee River and Pony Lake fulvic acids proceeds through reaction with 3DOM*;35 similar mechanisms were attributed to sulfonamide transformation in real river water and wastewater solutions with less than 7 (mg C) L−1. However, agricultural wastewater, lagoon water, and runoff all have high DOC concentrations (i.e., greater than 50 (mg C) L−1),36 and the role of agricultural DOM at these levels has not been characterized for photodegradation of antibiotics. Because DOM can also directly interact with antibiotics to form aqueous complexes,37−39 the photolytic fate of antibiotics in agriculturally relevant conditions is complex. Standardized DOM sources are not available for agricultural waste. For that reason, it is difficult to compare studies on the fate of antibiotics in animal waste as the variation in antibiotic photoreactivity with DOM from different sources is unknown. The objectives of this work were to (1) measure the generation of reactive species from poultry litter-derived DOM as a function of DOC concentration (up to 140 mg L−1) and (2) deconvolute the photolysis kinetics of antibiotics in simulated agriculturally impacted water matrices to determine the dominant reaction mechanisms. We focused on photodegradation of four antibiotics, namely CTC, ciprofloxacin (CIP), roxarsone (ROX), and sulfamethoxazole (SMX), in solutions containing DOM from three poultry litter sources or SRN. The selected antibiotics stem from four classes (i.e., tetracyclines, fluoroquinolones, organoarsenicals, and sulfonamides, respectively) used in the global poultry industry40−42 and previously detected in poultry litter.7,43−48 ROX and other organoarsenicals have been banned in Europe and North America,49,50 but these chemicals are still widely used around the world.51 CIP, CTC, and SMX have been identified as “critically important to human health” by the World Health Organization.42 An improved understanding of the photolytic fate of antibiotics in agriculture-impacted waters is necessary to mitigate environmental and human health concerns associated with antibiotic use in intensive agriculture.

2. MATERIALS AND METHODS 2.1. Chemicals and Reagents. CIP (>99%), CTC (>80%), ROX (>95%), and SMX (>95%) were purchased from SigmaAldrich (St. Louis, MO). The following scavengers and probe compounds were secured from Fisher Scientific (Pittsburgh, PA) for competition kinetics analysis: 2,4,6-trimethylphenol (TMP, probe for 3DOM*), furfuryl alcohol (FFA, probe for 1O2), parachlorobenzoic acid (pCBA, probe for HO•), sodium azide (NaN3, scavenger for HO• and 1O2), sorbic acid (scavenger for 3 DOM*), and tert-butanol (t-BuOH, scavenger for HO•). Buffers were constructed using formic acid and monobasic, dibasic, and tribasic sodium phosphate salts from Fisher Scientific. The Rose Bengal 1O2 sensitizer was obtained from Fisher Scientific. SRN was acquired from the International Humic Substances Society (Denver, CO). All solutions were generated in deionized (DI) water produced from an in-house system that employs sequential adsorption, ion exchange, reverse osmosis, and ultraviolet (UV) disinfection processes (Neu-Ion Systems; Baltimore, MD). 2.2. DOM Characterization. Poultry litter samples from three commercial farms in the Chesapeake Bay watershed were used as DOM sources. Litter samples were oven-dried at 40 °C, sieved (1.19 mm), and homogenized. Poultry litter extracts (PLEs) were prepared at room temperature by adding 40 g of the dried, homogenized poultry litter to 1 L of DI, shaking for 20 min at 250 rpm, and centrifuging at 14,000 g for 45 min at 26 °C. Aliquots (1 mL) of 0.45 μm filtered extracts were stored at −20 °C, and working solutions were diluted as needed. These extracts were constructed to simulate DOM leaching from poultry litter during irrigation and/or rainfall events. A standard SRN solution was made by adding 2.5 g of SRN to 1 L of DI. The resulting PLE and SRN stock solutions exhibited similar absorbance in the 310−410 nm range. Total organic carbon (TOC) concentrations were measured using a Shimadzu TOC-L instrument (Columbia, MD). Samples were typically diluted 100 times before measuring UV absorbance with (i) a 1 cm quartz cuvette in the Evolution 600 spectrophotometer (Thermo; Waltham, MA) or (ii) UVtransparent 96-well plates in the Eon microplate reader (BioTek; Winooski, VT). Fluorescence excitation−emission matrices (EEMs) were recorded for 100× diluted samples in 1 cm quartz cuvettes using a Cary Eclipse fluorescence spectrophotometer (Varian; Walnut Creek, CA). 2.3. Chemical Analysis. The concentrations of the four antibiotics and the three reactive species probe compounds were measured by liquid chromatography with diode array detection and tandem mass spectrometry (LC-DAD-MS/MS; Thermo UltiMate 3000 with Quantum Access Max). LC-MS grade water and methanol (both with 0.1% formic acid) were used to generate the mobile phase. The analytes were separated on a Waters Symmetry C18 column (2.1 × 150 mm, 3.5 μm). Details on the elution gradient, analyte retention times, DAD settings, and MS/MS parameters are provided in Figure S1, Table S1, and Text S1 of the Supporting Information (SI). 2.4. Photodegradation Experiments. Photodegradation experiments were conducted in a merry-go-round Rayonet RMR 600 reactor (Southern New England Ultraviolet Inc.; Branford, CT) equipped with bulbs emitting in the 310−410 nm range (see Figure S2 in the SI). The average incident photon flux of the system was calculated to be 2.51 × 10−5 Ein L−1 s−1 using the ferrioxalate actinometer52,53 with an average quantum yield of 1.23 mol Ein−1.54 Experimental solutions containing 1 mg L−1 12311

DOI: 10.1021/acs.est.7b03482 Environ. Sci. Technol. 2017, 51, 12310−12320

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Environmental Science & Technology Table 1. Properties of 100× Diluted PLE and SRN Stock Solutions

a

DOM

screening factor

E2/E3

DOC (mg L‑1)

UV254 (cm‑1)

SUVA254 (L (mg C)‑1 m‑1)

SRN PLE-0 PLE-1 PLE-2

0.794 ± 0.008 0.843 ± 0.008 0.847 ± 0.009 0.851 ± 0.009

4.72 ± 0.20 5.36 ± 0.99 5.74 ± 1.25 5.39 ± 0.93

9.20 ± 1.73 30.97 ± 2.34 28.32 ± 3.82 35.02 ± 6.42

0.491 ± 0.005 0.390 ± 0.004 0.384 ± 0.004 0.349 ± 0.004

5.84 ± 1.18 1.32 ± 0.12 1.48 ± 0.28 1.09 ± 0.24

a

The error values for all entries represent 95% confidence intervals obtained from triplicate measurements.

Figure 1. Steady state concentrations of (a, c) 1O2 and (b, d) 3DOM* produced during irradiation of solutions containing DOC from SRN, PLE-0, PLE1, and PLE-2 plotted as a function of (a, b) DOC concentration and (c, d) screening factor. Each symbol represents an average steady state concentration determined from six measurements of probe compounds over 90 min of irradiation. Error bars are 95% confidence intervals around the mean steady state concentration from time series experiments. The legend in (a) also applies to (b), (c), and (d).

antibiotic, variable DOC (from SRN or PLE stock solutions), and 10 mM phosphate buffer (pH 6.8, based on the natural pH of the PLEs) were added to 10 mL quartz tubes and irradiated. Aliquots of 100 μL were extracted at predetermined times and diluted to 1 mL with 0.1% formic acid prior to analysis by LCDAD-MS/MS. To determine steady state concentrations of reactive species, TMP, FFA, and pCBA were added at 25, 25, and 1 mg L−1, respectively. To selectively scavenge reactive species, NaN3, sorbic acid, and t-BuOH were added at concentrations of 100, 50, and 10 mg L−1, respectively. When determining the second-order reaction rate constants for antibiotics with 1O2, Rose Bengal was dosed at 25 mg L−1. The observed photodegradation kinetics for all antibiotics was fit to a time-based, pseudo-first-order model (eq 1).

ln

[AB]t ′ t = − kobs [AB]o

(1)

The observed time-based rate constants (kobs ′ ) for photodegradation of antibiotics were calculated as the slope of the natural logarithm of the normalized antibiotic concentration ([AB]t/[AB]o) plotted against time, t. The overall k′obs was comprised of contributions from direct photolysis and reactions with HO•, 1O2, and 3DOM*, as described in eq 2. ′ = Skd,AB ′ + k HO ″ •,AB[HO•]ss + k1″ kobs O

[1O2 ]ss + k 3″DOM*,AB[3DOM*]ss

2 ,AB

• + k′ 1 ′ + kobs,HO ′ = kobs,d + kobs, ′ 3 DOM * obs, O 2

(2)

In eq 2, S is the screening correction, kd,AB ′ is the pseudo-firstorder rate constant for direct photolysis of the antibiotic, ki,AB ″ is 12312


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Environmental Science & Technology

humic-like components.57 The composition of SRN agrees with the higher SUVA254 and lower E2/E3 ratios reported in Table 1 and indicates a fairly aromatic DOM with a higher molecular weight distribution compared to the poultry litter-derived DOM.58 The observed compositional differences suggest that DOM from poultry litter and SRN will exhibit variable generation of, and reaction with, reactive species during photolysis.59−61 The 1O2 concentrations were measured using the FFA probe compound (k1″O2,FFA = 1.0 × 108 M−1 s−162). For all DOM sources, steady state concentrations of 1O2 (see Figure 1a) were in the 10−13−10−12 M range, in agreement with previous reports for surface waters.29 For the PLEs, 1O2 concentrations were comparable for specific DOC levels, potentially due to the similar PLE compositions.56 Steady state concentrations of 1O2 in the SRN solutions were higher than those in the PLEcontaining solutions. For example, 37 (mg C) L−1 resulted in 1.19 × 10−12 M 1O2 for the SRN matrix, nearly double that of PLE-2 (i.e., 6.18 × 10−13 M 1O2). As predicted by previous studies,63 the steady state concentration of 1O2 increased linearly for all sources at low DOC levels (i.e., less than 30 mg L−1). Above this threshold, a noticeable change was observed for the 1 O2 yield. When these data were plotted against the screening factor, the steady state concentrations of 1O2 were similar for all four DOM sources (Figure 1c). For a screening factor of 0.95−0.96, the 1O2 concentrations were 1.96 × 10−13 M and 2.14 × 10−13 M for SRN and PLE-1, respectively. This observation suggests that the quantum yield for 1O2 formation was similar for SRN and the PLEs, and the differences in steady state 1O2 concentrations at higher DOM levels are associated with the specific rate of light absorption of the DOM sources. Given the lower SUVA254 and higher E2/E3 of the PLEs, higher 1O2 concentrations were expected based on previous research reporting higher 1O2 quantum yields for DOM with lower molecular weight distributions.64 In this case, the inherent differences in the DOM composition (e.g., the high fulvic- and humic-like character of SRN compared to the relatively high protein-like fluorescence of the PLEs) or coextracted inorganic ions may cause different trends in 1O2 quantum yield or scavenge some of the produced 1O2. The observed steady state concentrations of 3DOM* were determined using the TMP probe molecule (k3″DOM*,TMP = 3.0 × 109 M−1 s−165), while accounting for TMP reaction with 1O2 (k1″O2,TMP = 6.7 × 107 M−1 s−166). The TMP probe evaluates the electron transfer pathway associated with 3DOM* reaction, which is the expected mechanism for DOM reactivity with antibiotics. Like 1O2, 3DOM* concentrations increased in a linear fashion at low DOC concentrations; however, 3DOM* levels plateaued around 30 (mg C) L−1 for PLE-0 and PLE-2, presumably due to reaction with DOM. Wenk et al.67 also reported scavenging of triplet state sensitizers by aquatic and terrestrial DOM at 20−70 (mg C) L−1. The three PLEs generated different 3DOM* concentrations for the same DOC levels (Figure 1b) and screening factors (Figure 1d). For high DOM (i.e., > 30 (mg C) L−1), 3DOM* concentrations in the PLE-0 solution were approximately 50% of those for PLE-2, suggesting that formation of 3DOM* was slower for PLE-0 or the reaction of 3DOM* with DOM was faster for PLE-0. Steady state concentrations of 3DOM* for the SRN matrix were higher than those for the PLEs. These findings may be attributed to the greater presence of fulvic-like molecules in SRN, as these

the second-order rate constant for antibiotic reaction with reactive species i, and k′obs,i is the observed rate constant for direct photolysis or reaction with HO•, 1O2, or 3DOM*; the ss subscript indicates steady state concentration. In the absence of DOM, k′obs was interpreted as k′d,AB. The value for S was dependent on the DOM matrix (see Text S2 in the SI) and varied from 0.99 to 0.59 for solutions containing 2−140 (mg C) L−1 for the PLEs and 0.5−40 (mg C) L−1 for SRN. When available, second-order rate constants for reaction of antibiotics with reactive species were collected from the literature; otherwise, these rate constants were experimentally measured. Steady state concentrations of reactive species were obtained using the aforementioned probe molecules. Expressions for the mechanism-specific rate constants are provided in eqs 3−6 as a function of DOM concentration. Detailed derivations of these expressions are available in Text S3 (see the SI). ′ = kobs,d

Sk f,′3 AB * 3

*,DOM 1 + Pd,AB [DOM] 3 AB *

• = ′ kobs,HO

′ •[DOM] k f,HO •

HO ,DOM [AB] + PHO • ,AB [DOM] +

k H′ O,HO• 2

k HO ″ •,AB

(4)

k f,′1 O [DOM] 2

kobs, ′ 1O = 2

(3)

[AB] +

kobs, ′ 3 DOM * =

1

,DOM P 1OO2,AB [DOM] 2

+

k′

H2O,1O2

k1″

(5)

O2,AB

k f,′3 DOM *[DOM] 3

DOM*,DOM [AB] + P 3DOM [DOM] + *,AB

k′

H2O,3 DOM *

k 3″

DOM*,AB

(6)

In eqs 3−6, k′f,i is the rate constant for the reaction governing formation of excited antibiotics (3AB*) and reactive species (i.e., 3 *,DOM is the preference ratio for HO•, 1O2, and 3DOM*), Pd,AB 3 AB* 3 AB* quenching to ground state AB by DOM over direct ′ *,DOM/k′d,3AB*), Pi,DOM is the photolysis of 3AB* (defined as k3′AB i,AB preference ratio for reaction of reactive species with DOM over their reaction with an antibiotic (defined as ki,DOM ′′ /ki,AB ′′ ), and k′H2O,i is the rate constant for deactivation of a reactive species by water. DOM was quantified as DOC concentration. The overall photodegradation kinetics of individual antibiotics in solutions with DOM was fit to the expanded form of eq 2 (substitution of eqs 3−6) using a least-squares approach in OriginPro (OriginLab Corp.; Northampton, MA) and Microsoft Excel. Multiple starting points were employed for each antibiotic to ensure that the global minimum was reported and that the corresponding rate constants represented the best fit to experimental data.

3. RESULTS AND DISCUSSION 3.1. Generation of Reactive Species from Poultry Litter. Physicochemical properties of the DOM sources are reported in Table 1. The absorbance spectra for 310−410 nm were similar for the four DOM stock solutions (see Figure S2a in the SI), resulting in comparable screening corrections (see Table S2 in the SI). Fluorescence analysis (Figure S2c-e in the SI) of the PLEs revealed terrestrial humic-, microbial humic-, and proteinlike signatures.55,56 The fluorescence intensity of the protein-like pool in the PLEs was significantly greater than that of SRN (Figure S2b in the SI), which mostly consists of fulvic acid- and 12313


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Environmental Science & Technology Table 2. Physicochemical Properties of the Four Antibiotics of Concern

a

Dominant structure at pH 6.8. bpKa values were collected from ref 83 or ref 86. cAverage molar absorptivity for 310−410 nm at pH 6.8. dError values represent 95% confidence intervals. eFluence-based, pseudo-first-order rate constant for direct photolysis at 310−410 nm at pH 6.8.

Figure 2. Effect of DOM source and concentration on the observed transformation kinetics of (a) CIP, (b) ROX, (c) CTC, and (d) SMX. The undiluted DOC concentrations for SRN, PLE-0, PLE-1, and PLE-2 were 0.92, 3.10, 2.83, and, 3.50 g L−1, respectively. Columns represent the difference of the observed rate constant for antibiotic degradation in DI water (no DOM) with that from solutions with DOM. Error bars are 95% confidence intervals propagated from the mean rate constants from time series experiments. Positive and negative values indicate that photodegradation was sensitized and suppressed, respectively, by DOM. The legend in (a) applies to (b), (c), and (d).

No degradation of pCBA was observed in these experiments,

compounds have been associated with greater steady state concentrations of 3DOM*.68 A parallel factor analysis of EEMs for two of the PLEs used in this study did not identify fulvic acid components.56

suggesting that HO• was not generated at an appreciable concentration (i.e., [HO•]ss < 10−18 M for our experimental 12314


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Figure 3. Effects of SRN, PLE-0, PLE-1, and PLE-2 (columns i, ii, iii, and iv, respectively) on the photodegradation of CIP, ROX, CTC, and SMX (rows a, b, c, and d, respectively). Symbols are experimentally measured rate constants. Error bars are 95% confidence intervals around the mean rate constant from time series experiments. The curves stem from the expanded form of eq 2 (with rate constants from Table 3). The legend in (a) also applies to (b), (c), and (d).

conditions). Therefore, HO• reaction mechanisms in eq 2 were not considered during fitting of observed rate constants. 3.2. Direct Photolysis of Antibiotics. At pH 6.8, k′d,AB for the four antibiotics ranged over 3 orders of magnitude: CIP, 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, 1.33 (±0.27) × 10−6 s−1. The fluence-based, pseudo-first-order rate constant and average molar absorptivity at 310−410 nm for each antibiotic are provided in Table 2. More details on these calculations can be found in Text S4 (see the SI). The photoreactivity of the antibiotics did not directly correlate to their molar absorptivity. For example, CIP and ROX have similar absorbance in the 310− 410 nm range, but CIP is 3 orders of magnitude more reactive than ROX. These findings indicate that the quantum yields vary significantly between antibiotic classes. Self-sensitization of antibiotics during direct photolysis was not found to be significant (see Figure S3 in the SI). Specific reaction mechanisms for direct photolysis have been reported elsewhere for CIP,69,70 CTC,71 and SMX.72,73 The indirect photolysis of antibiotics due to irradiation of DOM from PLE-0, PLE-1, PLE2, and SRN at agriculturally relevant concentrations is discussed below. 3.3. Effect of DOM on Antibiotic Photodegradation. The net differences in the observed pseudo-first-order rate

constants for CIP, CTC, ROX, and SMX photodegradation in the presence/absence of DOM are shown in Figure 2. In Figure 2, positive and negative differences imply sensitization and inhibition, respectively, of antibiotic degradation by DOM. The observed rate constants for the four antibiotics showed different trends with DOC concentration. For instance, the observed rate constant for CIP consistently decreased with increasing DOC for all sources; however, ROX degradation was enhanced over the same DOC gradient. The observed rate constant for CTC degradation exhibited marginal changes for the examined DOC levels, although differences were apparent between DOM sources. Photodegradation of SMX was dependent on both DOM source and concentration. For example, the observed rate constant for SMX in the PLE-0 matrix increased steadily from 1000× to 100× dilution, before decreasing to the approximate magnitude for direct photolysis. For the SRN solution, observed rate constants for SMX degradation decreased steadily with DOC. These trends can be attributed to the dominant degradation mechanisms, which depend on antibiotic structure, and DOM generation of, and reaction with, reactive species. Dominant photodegradation mechanisms for the four antibiotics of interest were established by selective reactive species scavenging experiments performed in solutions containing 9.2 (mg C) L−1 SRN and 1 mg L−1 antibiotic. CIP photodegradation 12315


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Table 3. Rate Constants and Preference Ratios for Direct and Indirect Photolysis of the Four Antibiotics of Concern in the Four DOM Matrices Antibiotic-Specific Properties parameter

unit

CIP

ROX

CTC

SMX

k′f,3AB* k1″O2,AB

s−1 M−1 s−1

7.47 × 10−3 b

1.33 × 10−6 3.05 × 107a

2.83 × 10−4 1.50 × 106c

6.62 × 10−5 b

k3″DOM*,AB (SRN) k3″DOM*,AB (PLE-0) k″3DOM*,AB (PLE-1) k3″DOM*,AB (PLE-2)

M−1 s−1 M−1 s−1 M−1 s−1 M−1 s−1

b b b b

7.92 × 109 1.16 × 1010 1.46 × 109 5.74 × 109

7.18 × 107 2.91 × 109 1.14 × 108 1.04 × 109

parameter

SRN

PLE-0

PLE-1

PLE-2

k′f,1O2

s−1

2.45 × 10−4

3.21 × 10−6

1.11 × 10−5

2.66 × 10−6

k′f,3DOM* k1″O2,DOM

s−1 Mc−1 s−1

2.24 × 10−4 1.80 × 109

9.28 × 10−4 1.64 × 107

7.96 × 10−4 5.36 × 107

5.09 × 10−3 1.01 × 107

k″3DOM*,DOM

Mc−1 s−1

6.19 × 109

1.46 × 108

2.35 × 108

parameter

unit

SRN

PLE-0

PLE-1

PLE-2

Mc−1

3.22 × 103

1.83 × 103

1.07 × 103

1.28 × 103

Mc−1 M Mc−1

2.24 × 106 5.90 × 101

2.22 × 103 5.93 × 10−1

2.17 × 102 1.76 × 10°

6.54 × 104 3.32 × 10−1

Mc−1 M Mc−1

1.29 × 102 2.98 × 102

3.30 × 103 6.81 × 102

8.84 × 103 1.99 × 102

8.17 × 102 2.43 × 102

M Mc−1

7.81 × 10−1

6.23 × 10−2

4.12 × 10−1

4.09 × 10−2

Mc−1 M Mc−1

5.46 × 104 8.52 × 101

6.70 × 104 2.49 × 10−1

1.50 × 104 1.28 × 10°

1.54 × 104 2.27 × 10−1

CIP 3 *,DOM Pd,AB 3 AB* ROX 3 *,DOM Pd,AB 3 AB* 1 O2,DOM P1O2,AB CTC 3 *,DOM Pd,AB 3 AB* 1 O2,DOM P1O2,AB 3

*,DOM P3DOM DOM*,AB SMX 3 *,DOM Pd,AB 3 AB* 3 *,DOM P3DOM DOM*,AB

a

unit

b b b b DOM-Specific Properties

7.23 × 108 Preference Ratios

See Text S6 in the SI. bNegligible. cFrom ref 78.

demonstrated faster quenching of 3CIP* to CIP. Porras et al.75 reported that CIP photolysis was inhibited in the presence of natural organic matter from a Nordic reservoir. This report, in combination with our findings, suggests that DOM composition plays a critical role in the fate of the 3CIP* species. Even though CIP photoreactivity was suppressed by DOM through screening effects and quenching of 3CIP* to CIP, the k′obs in the current study remained in the 10−4 s−1 range, even at the highest DOC concentrations. This observed rate constant was 1−2 orders of magnitude higher than those calculated for other antibiotics in this study. These findings demonstrate that CIP is effectively photodegraded in natural sunlight, even in agriculturally impacted waters with high DOC. 3.3.2. Roxarsone − Reaction with 1O2. Selective reactive species scavenging experiments indicated that ROX reaction with 1 O2 was the dominant photodegradation mechanism (see Text S5 and Figure S7 in the SI). Eq 3 and eq 5 were, therefore, used to explain the observed photodegradation trends. The second-order rate constant for reaction of ROX with 1O2 was calculated to be 3.05 (±0.36) × 107 M−1 s−1 (see Text S6 in the SI). The magnitude of this rate constant, along with the measured steady state concentrations, confirmed the importance of the 1O2 mechanism for ROX degradation. Figure 3b shows the observed rate constant for ROX photodegradation in solutions with different DOM sources. In general, the ROX transformation kinetics increased with DOC concentration for all sources, although some inhibitory effects were observed at low concentrations of PLE-0 and PLE-2 due to

was not significantly enhanced by DOM, but the other antibiotics were sensitized as follows: ROX, 1O2; SMX, 3DOM*; and, CTC, 3 DOM* and 1O2 (see Figure S4 in the SI). The following subsections expand on the photoreactivity of individual antibiotics in the various DOM matrices. 3.3.1. Ciprofloxacin − Screening and Quenching of 3CIP* to CIP by DOM. DOM, regardless of source, inhibited CIP transformation. Previous studies have reported suppression of fluoroquinolone photolysis in surface water and wastewater effluent26,69 due to screening effects. Given the low reactivity of CIP with 1O2 and 3DOM*,74 direct photolysis was the dominant photodegradation mechanism. The time-based degradation trends for CIP in solutions containing 3−123 (mg C) L−1 from PLE-0 are provided in Figure S5 (see the SI). Figure 3a shows the observed rate constant for CIP plotted against DOC concentration for all DOM sources. The low k′obs for CIP at higher DOC concentrations was not accounted for by screening corrections alone (see Figure S6 in the SI). These results suggest that DOM quenched 3CIP* to ground state CIP, as indicated by Reaction S4 in Text S3 of the SI. Similar quenching mechanisms were reported by Wenk et al.67 for solutions containing 22−72 (mg C) L−1 from various DOM sources. The observed photolysis kinetics of CIP were fit to eq 3. The 3 *,DOM preference ratio Pd,CIP is reported in Table 3 for the four 3 CIP* DOM sources. These values were comparable for all three PLEs, indicating that the reaction of poultry litter-derived DOM with 3 CIP* is fairly consistent. This result likely stems from the similar composition of the three PLEs. The more aromatic SRN matrix 12316


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finding supports sensitization of SMX at high DOC concentrations. The k′3DOM ′ *,SMX values differed for each PLE source and were determined to be 2.9 × 109, 1.1 × 108, and 1.0 × 109 M−1 s−1 for PLE-0, PLE-1, and PLE-2, respectively. The enhanced photoreactivity of SMX in the PLEs may be associated with the microbial humic-like components. Previous studies have indicated that 3DOM* from autochthonous sources, including wastewater effluent and lake water, enhanced the photodegradation of sulfa-drugs35,80 compared to 3DOM* from Suwannee River. This effect may be associated with greater antioxidant properties of SRN or greater oxidation potential of nonterrestrial sources.65 3.4. Environmental Significance. In this study, direct and indirect photolysis kinetics and mechanisms have been analyzed for four antibiotics at agriculturally relevant conditions using organic matter extracted from three poultry litters. While previous studies have described indirect photolysis of contaminants of emerging concern in surface waters with low DOC, this study shows that antibiotic photodegradation also occurs at the high DOM levels characteristic of agriculturally impacted waters. The calculated parameters associated with formation and quenching of reactive species in the PLE and SRN solutions indicate that agriculturally derived DOM does not impact photodegradation in the same way as common surrogate materials like SRN. These results suggest that the photochemistry of antibiotics in surface water significantly differs from agricultural waters. As more than 70% of antibiotics sold in the United States are used in agriculture,81 the fundamental photodegradation mechanisms and kinetics reported in this study constitute important advances in understanding antibiotic fate in the environment. To gain further insight into the photodegradation of antibiotics in agricultural systems, future studies should investigate real agriculturally impacted waters with mixed DOM sources. The observed photoreactivity varied for each antibiotic; furthermore, the impacts of DOM source and concentration manifested in different ways for each antibiotic. Degradation of CIP was inhibited due to screening effects and quenching of the

screening and DOM quenching of 3ROX* to ROX. Quenching of 3ROX* to ROX varied with DOM source, as indicated by the 3 *,DOM calculated values for Pd,ROX : 2.2 × 106 Mc−1 for SRN; 2.2 × 3 ROX* 3 2 −1 10 Mc for PLE-0; 2.2 × 10 Mc−1 for PLE-1; and, 6.5 × 104 Mc−1 for PLE-2. These findings verified the different behaviors of the DOM sources at low DOC levels. At high DOM concentrations, the overall contribution of the 1 O2 reaction increased (see Figure S8 in SI). As expected from Figure 1a, the rate constant for formation of 1O2 was higher for SRN (kf,′1O2 = 2.5 × 10−4 s−1) compared to the PLEs (kf,′1O2 = 2.7− 11 × 10−6 s−1). The rate constants for reaction of bulk DOM with 1

O2 were source-dependent. For instance, the P1OO22,DOM ,ROX ratio for ROX varied in the range of 0.3−59 M Mc−1, indicating that DOM differentially competed with ROX for reaction with 1O2. 3.3.3. Chlortetracycline − Reaction with 1O2 and 3DOM*. Self-sensitized photodegradation of tetracyclines by 1O2 has been reported,76 but this mechanism was not found to be significant for the experimental conditions examined here (see Text S3 and Figure S3 in the SI). Reactive species scavenging studies (see Text S5 and Figure S7 in the SI) indicated that CTC reactions with 3DOM* and 1O2 were important in the presence of 9.2 (mg C) L−1 of SRN. Previous authors have described the role of 1O2 in CTC phototransformation.77,78 Figure 3c shows the variation of kobs ′ for CTC photodegradation as a function of DOM source and concentration. The observed rate constant was fit using eq 3, eq 5, and eq 6. The overall contribution of 1O2 to CTC degradation was low (see Figure S9 in the SI) due to slow reaction kinetics (see Table 3). The net inhibition of CTC degradation in solutions containing PLE-1 at DOC concentrations less than 20 mg L−1 was mostly associated with light screening and quenching of 3 CTC* to CTC by DOM. At higher DOM concentrations, reaction with 3DOM* dominated the overall degradation of 3 *,DOM CTC. The preference ratio, P3DOM DOM*,CTC , was higher for SRN than 1

the PLEs, indicating that 3DOM* originating from PLE and SRN showed source-dependent reactivity with DOM and CTC. The second-order rate constants obtained for reaction of CTC with

CIP* intermediate to ground state CIP by DOM. CTC, ROX, and SMX showed varying degrees of sensitization due to selective reactivity with 3DOM* and 1O2. Previous studies have reported phototransformation of antibiotics in engineered systems82−85 causes formation of antimicrobially active transformation products. The effects of agricultural DOM on these reactions, namely the differences in reaction pathway for direct photolysis and 1O2 and 3DOM* mediated processes, are a critical area for future study. Given the global use of diverse antibiotics to raise food-producing animals, the results of this study highlight the need to determine the photolytic fate of antibiotics in waste management systems, such as anaerobic lagoons, and agricultural runoff. Findings from this study also have implications for the impact of runoff from animal-derived waste streams on surface water, wherein mixed DOM sources may differentially affect the fate of antibiotics and other contaminants of emerging concern, including hormones, pesticides, and herbicides. 3

DOM* from each source were as follows: 1.5 × 109 Mc−1 s−1 for PLE-1; 5.7 × 109 Mc−1 s−1 for PLE-2; 7.9 × 109 Mc−1 s−1 for SRN; and, 1.2 × 1010 Mc−1 s−1 for PLE-0. 3.3.4. Sulfamethoxazole − Reaction with 3DOM*. Even though SMX does not absorb much light above 325 nm, the contribution of direct photolysis is significant due to the high quantum yield (i.e., 0.50 ± 0.09 at pH 5.3 for solar irradiation79). Boreen et al.79 reported that SMX does not react with 1O2. Therefore, any enhancement in the photodegradation rate of SMX in DOM-containing solutions is expected to occur through 3

DOM* mechanisms. Similar conclusions have been reported for photodegradation of SMX in wastewater.35,80 Figure 3d shows the effect of DOM on the observed rate constant for SMX, which was fit using eq 3 and eq 6. DOM content and type had complex effects on the observed SMX rate constant. For solutions containing SRN, k obs ′ consistently decreased with increasing DOC due to the low ′ *,SMX (7.2 × 107 M−1 s−1). At low DOM levels, 3SMX* was k3′DOM quenched to ground state SMX by DOM. Although SMX showed low reactivity with 3DOM* from SRN, reaction with 3

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ASSOCIATED CONTENT

S Supporting Information *

DOM* was dominant over direct photolysis at high DOM concentrations for all sources (see Figure S10 in the SI). This 3

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.7b03482. 12317


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Method details; analysis and fitting of kinetic model; determination of second-order rate constant for roxarsone with singlet oxygen; supporting tables and figures (PDF)

(15) Dolliver, H.; Kumar, K.; Gupta, S. Sulfamethazine uptake by plants from manure-amended soil. J. Environ. Qual. 2007, 36 (4), 1224− 1230. (16) Kumar, K.; Gupta, S.; Baidoo, S.; Chander, Y.; Rosen, C. Antibiotic uptake by plants from soil fertilized with animal manure. J. Environ. Qual. 2005, 34 (6), 2082−2085. (17) Broekaert, N.; Daeseleire, E.; Delezie, E.; Vandecasteele, B.; De Beer, T.; Van Poucke, C. Can the use of coccidiostats in poultry breeding lead to residues in vegetables? An experimental study. J. Agric. Food Chem. 2012, 60 (50), 12411−12418. (18) Yao, L.; Li, G.; Dang, Z.; Yang, B.; He, Z.; Zhou, C. Uptake and transport of roxarsone and its metabolites in water spinach as affected by phosphate supply. Environ. Toxicol. Chem. 2010, 29 (4), 947−951. (19) Colomer-Lluch, M.; Imamovic, L.; Jofre, J.; Muniesa, M. Bacteriophages carrying antibiotic resistance genes in fecal waste from cattle, pigs, and poultry. Antimicrob. Agents Chemother. 2011, 55 (10), 4908−4911. (20) Himathongkham, S.; Riemann, H.; Bahari, S.; Nuanualsuwan, S.; Kass, P.; Cliver, D. Survival of Salmonella typhimurium and Escherichia coli O157: H7 in poultry manure and manure slurry at sublethal temperatures. Avian Dis. 2000, 44 (4), 853−860. (21) Hayes, J. R.; English, L. L.; Carr, L. E.; Wagner, D. D.; Joseph, S. W. Multiple-antibiotic resistance of Enterococcus spp. isolated from commercial poultry production environments. Appl. Environ. Microbiol. 2004, 70 (10), 6005−6011. (22) Boreen, A. L.; Arnold, W. A.; McNeill, K. Photodegradation of pharmaceuticals in the aquatic environment: A review. Aquat. Sci. 2003, 65 (4), 320−341. (23) OECD, OECD Guidelines for the Testing of Chemicals: Guideline 316; Adopted: October 3, 2008; DOI: 10.1787/9789264067585-en. (24) Batchu, S. R.; Panditi, V. R.; Gardinali, P. R. Photodegradation of sulfonamide antibiotics in simulated and natural sunlight: Implications for their environmental fate. J. Environ. Sci. Health, Part B 2014, 49 (3), 200−211. (25) Xu, H.; Cooper, W. J.; Jung, J.; Song, W. Photosensitized degradation of amoxicillin in natural organic matter isolate solutions. Water Res. 2011, 45 (2), 632−638. (26) Wammer, K. H.; Korte, A. R.; Lundeen, R. A.; Sundberg, J. E.; McNeill, K.; Arnold, W. A. Direct photochemistry of three fluoroquinolone antibacterials: Norfloxacin, ofloxacin, and enrofloxacin. Water Res. 2013, 47 (1), 439−448. (27) Guerard, J. J.; Miller, P. L.; Trouts, T. D.; Chin, Y.-P. The role of fulvic acid composition in the photosensitized degradation of aquatic contaminants. Aquat. Sci. 2009, 71 (2), 160−169. (28) Rosario-Ortiz, F. L.; Canonica, S. Probe compounds to assess the photochemical activity of dissolved organic matter. Environ. Sci. Technol. 2016, 50 (23), 12532−12547. (29) Burns, J. M.; Cooper, W. J.; Ferry, J. L.; King, D. W.; DiMento, B. P.; McNeill, K.; Miller, C. J.; Miller, W. L.; Peake, B. M.; Rusak, S. A. Methods for reactive oxygen species (ROS) detection in aqueous environments. Aquat. Sci. 2012, 74 (4), 683−734. (30) Lundeen, R. A.; Chu, C.; Sander, M.; McNeill, K. Photooxidation of the antimicrobial, nonribosomal peptide bacitracin A by singlet oxygen under environmentally relevant conditions. Environ. Sci. Technol. 2016, 50 (16), 8586−8595. (31) Wang, L.; Xu, H.; Cooper, W. J.; Song, W. Photochemical fate of beta-blockers in NOM enriched waters. Sci. Total Environ. 2012, 426, 289−295. (32) He, K.; Soares, A. D.; Adejumo, H.; McDiarmid, M.; Squibb, K.; Blaney, L. Detection of a wide variety of human and veterinary fluoroquinolone antibiotics in municipal wastewater and wastewaterimpacted surface water. J. Pharm. Biomed. Anal. 2015, 106, 136−143. (33) Mangalgiri, K. P.; He, K.; Blaney, L. Emerging contaminants: A potential human health concern for sensitive populations. PDA J. Pharm. Sci. Technol. 2015, 69 (2), 215−218. (34) Yan, S.; Song, W. Photo-transformation of pharmaceutically active compounds in the aqueous environment: A review. Environmental Science: Processes & Impacts 2014, 16 (4), 697−720.

AUTHOR INFORMATION

Corresponding Author

*Phone: +1-410-455-8608. Fax: +1-410-455-1049. E-mail: [email protected]. ORCID

Lee Blaney: 0000-0003-0181-1326 Notes

The authors declare no competing financial interest.

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

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REFERENCES

(1) Mangalgiri, K. P.; Adak, A.; Blaney, L. Organoarsenicals in poultry litter: Detection, fate, and toxicity. Environ. Int. 2015, 75, 68−80. (2) Campagnolo, E. R.; Johnson, K. R.; Karpati, A.; Rubin, C. S.; Kolpin, D. W.; Meyer, M. T.; Esteban, J. E.; Currier, R. W.; Smith, K.; Thu, K. M. Antimicrobial residues in animal waste and water resources proximal to large-scale swine and poultry feeding operations. Sci. Total Environ. 2002, 299 (1), 89−95. (3) Webb, K.; Fontenot, J. Medicinal drug residues in broiler litter and tissues from cattle fed litter. J. Anim. Sci. 1975, 41 (4), 1212−1217. (4) Furtula, V.; Farrell, E.; Diarrassouba, F.; Rempel, H.; Pritchard, J.; Diarra, M. Veterinary pharmaceuticals and antibiotic resistance of Escherichia coli isolates in poultry litter from commercial farms and controlled feeding trials. Poult. Sci. 2010, 89 (1), 180−188. (5) Van Epps, A.; Blaney, L. Antibiotic Residues in Animal Waste: Occurrence and Degradation in Conventional Agricultural Waste Management Practices. Curr. Pollution Rep 2016, 2 (3), 135−155. (6) Morrison, J. L. Distribution of arsenic from poultry litter in broiler chickens, soil, and crops. J. Agric. Food Chem. 1969, 17 (6), 1288−90. (7) Zhao, L.; Dong, Y. H.; Wang, H. Residues of veterinary antibiotics in manures from feedlot livestock in eight provinces of China. Sci. Total Environ. 2010, 408 (5), 1069−1075. (8) Ho, Y.; Zakaria, M. P.; Latif, P. A.; Saari, N. Simultaneous determination of veterinary antibiotics and hormone in broiler manure, soil and manure compost by liquid chromatography−tandem mass spectrometry. Journal of Chromatography A 2012, 1262, 160−168. (9) Kelleher, B.; Leahy, J.; Henihan, A.; O’dwyer, T.; Sutton, D.; Leahy, M. Advances in poultry litter disposal technology − A review. Bioresour. Technol. 2002, 83 (1), 27−36. (10) Gaskin, J. W.; Harris, G. H.; Franzluebbers, A.; Andrae, J. Poultry litter application on pastures and hayfields; Cooperative Extension document B1330; University of Georgia, College of Agricultural and Environmental Sciences: April 9, 2010. (11) Jiang, Z.; Li, P.; Wang, Y.; Li, B.; Wang, Y. Effects of roxarsone on the functional diversity of soil microbial community. Int. Biodeterior. Biodegrad. 2013, 76, 32−35. (12) Majalija, S.; Oweka, F.; Wito, G.; Lubowa, M.; Vudriko, P.; Nakamaya, F. Antibiotic susceptibility profiles of faecal Escherichia coli isolates from dip-litter broiler chickens in Northern and Central Uganda. Veterinary Research 2010, 3 (4), 75−80. (13) Reichel, R.; Patzelt, D.; Barleben, C.; Rosendahl, I.; Ellerbrock, R. H.; Thiele-Bruhn, S. Soil microbial community responses to sulfadiazine-contaminated manure in different soil microhabitats. Applied Soil Ecology 2014, 80, 15−25. (14) Covey, A.; Furbish, D.; Savage, K. Earthworms as agents for arsenic transport and transformation in roxarsone-impacted soil mesocosms: A μXANES and modeling study. Geoderma 2010, 156 (3), 99−111. 12318


Article

Environmental Science & Technology (35) Bahnmüller, S.; von Gunten, U.; Canonica, S. Sunlight-induced transformation of sulfadiazine and sulfamethoxazole in surface waters and wastewater effluents. Water Res. 2014, 57, 183−192. (36) Royer, I.; Angers, D. A.; Chantigny, M. H.; Simard, R. R.; Cluis, D. Dissolved organic carbon in runoff and tile-drain water under corn and forage fertilized with hog manure. J. Environ. Qual. 2007, 36 (3), 855− 863. (37) Fu, Q.-L.; He, J.-Z.; Blaney, L.; Zhou, D.-M. Roxarsone binding to soil-derived dissolved organic matter: Insights from multi-spectroscopic techniques. Chemosphere 2016, 155, 225−233. (38) Chen, Z.; Zhang, Y.; Gao, Y.; Boyd, S. A.; Zhu, D.; Li, H. Influence of dissolved organic matter on tetracycline bioavailability to an antibiotic-resistant bacterium. Environ. Sci. Technol. 2015, 49 (18), 10903−10910. (39) Pan, B.; Qiu, M.; Wu, M.; Zhang, D.; Peng, H.; Wu, D.; Xing, B. The opposite impacts of Cu and Mg cations on dissolved organic matter-ofloxacin interaction. Environ. Pollut. 2012, 161, 76−82. (40) FDA, FDA Approved Animal Drug Products, December 2016; US Food and Drug Administration: Silver Spring, MD, 2016. (41) Krishnasamy, V.; Otte, J.; Silbergeld, E. Antimicrobial use in Chinese swine and broiler poultry production. Antimicrobial Resistance and Infection Control 2015, 4 (1), 17. (42) WHO, Critically Important Antimicrobials for Human Medicine, 4th revision; WHO Document Production Services: Geneva, 2013. (43) Hu, X.; Luo, Y.; Zhou, Q. Simultaneous analysis of selected typical antibiotics in manure by microwave-assisted extraction and LC−MSn. Chromatographia 2010, 71 (3−4), 217−223. (44) Karcı, A.; Balcıoğlu, I. A. Investigation of the tetracycline, sulfonamide, and fluoroquinolone antimicrobial compounds in animal manure and agricultural soils in Turkey. Sci. Total Environ. 2009, 407 (16), 4652−4664. (45) Motoyama, M.; Nakagawa, S.; Tanoue, R.; Sato, Y.; Nomiyama, K.; Shinohara, R. Residues of pharmaceutical products in recycled organic manure produced from sewage sludge and solid waste from livestock and relationship to their fermentation level. Chemosphere 2011, 84 (4), 432−438. (46) Hancock, T.; Denver, J.; Riedel, G.; Miller, C. In Source, transport, and fate of arsenic in the Pocomoke River Basin, a poultry dominated Chesapeake Bay Watershed, Proceedings of Arsenic in the Environment Workshop. US Geological Survey. Open-File Report, 2001; 2001. (47) Brown, B.; Slaughter, A.; Schreiber, M. Controls on roxarsone transport in agricultural watersheds. Appl. Geochem. 2005, 20 (1), 123− 133. (48) Leal, R. M. P.; Figueira, R. F.; Tornisielo, V. L.; Regitano, J. B. Occurrence and sorption of fluoroquinolones in poultry litters and soils from São Paulo State, Brazil. Sci. Total Environ. 2012, 432, 344−349. (49) FDA Arsenic-based animal drugs and poultry. Available at < http://www.fda.gov/AnimalVeterinary/SafetyHealth/ ProductSafetyInformation/ucm257540.htm> (accessed Dec 12, 2016). (50) Lasky, T. Arsenic in chicken: A tale of data and policy. Journal of Epidemiology and Community Health 2017, 71 (1), 1−3. (51) Liu, X.; Zhang, W.; Hu, Y.; Cheng, H. Extraction and detection of organoarsenic feed additives and common arsenic species in environmental matrices by HPLC−ICP-MS. Microchem. J. 2013, 108, 38−45. (52) Hatchard, C.; Parker, C. A new sensitive chemical actinometer. II. Potassium ferrioxalate as a standard chemical actinometer. Proc. R. Soc. London, Ser. A 1956, 235, 518−536. (53) Leifer, A. The Kinetics of Environmental Aquatic Photochemistry; ACS Professional and Reference Book, American Chemical Society: Washington, DC, 1988. (54) Murov, S. L.; Carmichael, I.; Hug, G. L. Handbook of Photochemistry; CRC Press: Boca Raton, FL, 1993. (55) Coble, P. G. Marine optical biogeochemistry: the chemistry of ocean color. Chem. Rev. 2007, 107 (2), 402−418. (56) Mangalgiri, K. P.; Timko, S. A.; Gonsior, M.; Blaney, L. PARAFAC modeling of irradiation- and oxidation-induced changes in fluorescent dissolved organic matter extracted from poultry litter. Environ. Sci. Technol. 2017, 51 (14), 8036−8047.

(57) Timko, S. A.; Gonsior, M.; Cooper, W. J. Influence of pH on fluorescent dissolved organic matter photo-degradation. Water Res. 2015, 85, 266−274. (58) Helms, J. R.; Stubbins, A.; Ritchie, J. D.; Minor, E. C.; Kieber, D. J.; Mopper, K. Absorption spectral slopes and slope ratios as indicators of molecular weight, source, and photobleaching of chromophoric dissolved organic matter. Limnol. Oceanogr. 2008, 53 (3), 955−969. (59) Cory, R. M.; Cotner, J. B.; McNeill, K. Quantifying interactions between singlet oxygen and aquatic fulvic acids. Environ. Sci. Technol. 2009, 43 (3), 718−723. (60) Paul, A.; Hackbarth, S.; Vogt, R. D.; Röder, B.; Burnison, B. K.; Steinberg, C. E. Photogeneration of singlet oxygen by humic substances: comparison of humic substances of aquatic and terrestrial origin. Photochemical & Photobiological Sciences 2004, 3 (3), 273−280. (61) Alberts, J. J.; Takács, M. Total luminescence spectra of IHSS standard and reference fulvic acids, humic acids and natural organic matter: Comparison of aquatic and terrestrial source terms. Org. Geochem. 2004, 35 (3), 243−256. (62) Appiani, E.; Ossola, R.; Latch, D. E.; Erickson, P. R.; McNeill, K. Aqueous singlet oxygen reaction kinetics of furfuryl alcohol: effect of temperature, pH, and salt content. Environmental Science: Processes & Impacts 2017, 19 (4), 507−516. (63) Janssen, E. M.-L.; Erickson, P. R.; McNeill, K. Dual roles of dissolved organic matter as sensitizer and quencher in the photooxidation of tryptophan. Environ. Sci. Technol. 2014, 48 (9), 4916−4924. (64) Maizel, A. C.; Remucal, C. K. Molecular composition and photochemical reactivity of size-fractionated dissolved organic matter. Environ. Sci. Technol. 2017, 51 (4), 2113−2123. (65) McNeill, K.; Canonica, S. Triplet state dissolved organic matter in aquatic photochemistry: reaction mechanisms, substrate scope, and photophysical properties. Environmental Science: Processes & Impacts 2016, 18 (11), 1381−1399. (66) Tratnyek, P. G.; Hoigné, J. Photo-oxidation of 2,4,6trimethylphenol in aqueous laboratory solutions and natural waters: Kinetics of reaction with singlet oxygen. J. Photochem. Photobiol., A 1994, 84 (2), 153−160. (67) Wenk, J.; Eustis, S. N.; McNeill, K.; Canonica, S. Quenching of excited triplet states by dissolved natural organic matter. Environ. Sci. Technol. 2013, 47 (22), 12802−12810. (68) Rering, C.; Williams, K.; Hengel, M.; Tjeerdema, R. Comparison of direct and indirect photolysis in imazosulfuron photodegradation. J. Agric. Food Chem. 2017, 65 (15), 3103−3108. (69) Babić, S.; Periša, M.; Škorić, I. Photolytic degradation of norfloxacin, enrofloxacin and ciprofloxacin in various aqueous media. Chemosphere 2013, 91 (11), 1635−1642. (70) Wei, X.; Chen, J.; Xie, Q.; Zhang, S.; Ge, L.; Qiao, X. Distinct photolytic mechanisms and products for different dissociation species of ciprofloxacin. Environ. Sci. Technol. 2013, 47 (9), 4284−4290. (71) Chen, Y.; Li, H.; Wang, Z.; Tao, T.; Wei, D.; Hu, C. Photolysis of chlortetracycline in aqueous solution: Kinetics, toxicity and products. J. Environ. Sci. 2012, 24 (2), 254−260. (72) Niu, J.; Zhang, L.; Li, Y.; Zhao, J.; Lv, S.; Xiao, K. Effects of environmental factors on sulfamethoxazole photodegradation under simulated sunlight irradiation: kinetics and mechanism. J. Environ. Sci. 2013, 25 (6), 1098−1106. (73) Trovó, A. G.; Nogueira, R. F.; Agüera, A.; Sirtori, C.; FernándezAlba, A. R. Photodegradation of sulfamethoxazole in various aqueous media: Persistence, toxicity and photoproducts assessment. Chemosphere 2009, 77 (10), 1292−1298. (74) Martinez, L. J.; Sik, R. H.; Chignell, C. F. Fluoroquinolone antimicrobials: Singlet oxygen, superoxide and phototoxicity. Photochem. Photobiol. 1998, 67 (4), 399−403. (75) Porras, J.; Bedoya, C.; Silva-Agredo, J.; Santamaría, A.; Fernández, J. J.; Torres-Palma, R. A. Role of humic substances in the degradation pathways and residual antibacterial activity during the photodecomposition of the antibiotic ciprofloxacin in water. Water Res. 2016, 94, 1−9. (76) Chen, Y.; Li, H.; Wang, Z.; Tao, T.; Hu, C. Photoproducts of tetracycline and oxytetracycline involving self-sensitized oxidation in 12319


Article

Environmental Science & Technology aqueous solutions: Effects of Ca2+ and Mg2+. J. Environ. Sci. 2011, 23 (10), 1634−1639. (77) Salazar-Rábago, J.; Sánchez-Polo, M.; Rivera-Utrilla, J.; LeyvaRamos, R.; Ocampo-Pérez, R. Role of 1[O2]* in chlortetracycline degradation by solar radiation assisted by ruthenium metal complexes. Chem. Eng. J. 2016, 284, 896−904. (78) Castillo, C.; Criado, S.; Díaz, M.; García, N. A. Riboflavin as a sensitiser in the photodegradation of tetracyclines. Kinetics, mechanism and microbiological implications. Dyes Pigm. 2007, 72 (2), 178−184. (79) Boreen, A. L.; Arnold, W. A.; McNeill, K. Photochemical fate of sulfa drugs in the aquatic environment: Sulfa drugs containing fivemembered heterocyclic groups. Environ. Sci. Technol. 2004, 38 (14), 3933−3940. (80) Ryan, C. C.; Tan, D. T.; Arnold, W. A. Direct and indirect photolysis of sulfamethoxazole and trimethoprim in wastewater treatment plant effluent. Water Res. 2011, 45 (3), 1280−1286. (81) Landers, T. F.; Cohen, B.; Wittum, T. E.; Larson, E. L. A review of antibiotic use in food animals: Perspective, policy, and potential. Public Health Rep. 2012, 127 (1), 4−22. (82) Snowberger, S.; Adejumo, H.; He, K.; Mangalgiri, K. P.; Hopanna, M.; Soares, A. D.; Blaney, L. Direct photolysis of fluoroquinolone antibiotics at 253.7 nm: Specific reaction kinetics and formation of equally potent fluoroquinolone antibiotics. Environ. Sci. Technol. 2016, 50 (17), 9533−9542. (83) Adak, A.; Mangalgiri, K. P.; Lee, J.; Blaney, L. UV irradiation and UV-H2O2 advanced oxidation of the roxarsone and nitarsone organoarsenicals. Water Res. 2015, 70, 74−85. (84) Keen, O. S.; Linden, K. G. Degradation of antibiotic activity during UV/H2O2 advanced oxidation and photolysis in wastewater effluent. Environ. Sci. Technol. 2013, 47 (22), 13020−13030. (85) Pereira, V. J.; Linden, K. G.; Weinberg, H. S. Evaluation of UV irradiation for photolytic and oxidative degradation of pharmaceutical compounds in water. Water Res. 2007, 41 (19), 4413−4423. (86) Qiang, Z.; Adams, C. Potentiometric determination of acid dissociation constants (pKa) for human and veterinary antibiotics. Water Res. 2004, 38 (12), 2874−90.

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