Photodegradation of Veterinary Ionophore Antibiotics under UV and

Oct 24, 2014 - the photodegradation of three commonly used IPAs, monensin. (MON), salinomycin (SAL) and narasin (NAR), under UV and solar irradiation...
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Photodegradation of Veterinary Ionophore Antibiotics under UV and Solar Irradiation Peizhe Sun, Spyros G. Pavlostathis, and Ching-Hua Huang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/es5034525 • Publication Date (Web): 24 Oct 2014 Downloaded from http://pubs.acs.org on October 28, 2014

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Photodegradation of Veterinary Ionophore Antibiotics

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under UV and Solar Irradiation

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

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

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Georgia 30332, United States

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

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E-mail: [email protected]

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Revised manuscript submitted to

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Environmental Science & Technology

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October 24, 2014

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ABSTRACT

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

19

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

20

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

21

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

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

23

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

24

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

25

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

26

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

27

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

28

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

29

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

30

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

31

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

32

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

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

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showed that photodegradation of SAL eliminated its antibiotic properties against Bacillus

35

subtilis.

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INTRODUCTION

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

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

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

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

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

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

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

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protozoa.4-10

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

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

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

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

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

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

51

19

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

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

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

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occasionally in wastewater treatment plant effluents.17

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

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

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

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

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

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

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

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transformation

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

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

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

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

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

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

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carbonyl moiety in their structures.26

mechanisms

has

been

investigated

previously.

For

example,

70

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

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

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environmental conditions, measure the quantum yields, elucidate the transformation

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

R

HO

1

4 2

3

A O

OH

7 8

15

37

32

6

O

9

O

10

11

14 12

16

B 17

38 O

C 21

13 O 18

30

74 75 76 77

33

E

D 24

41 28 40

25 O 29

O

20 OH

OH

27

39

42

19

36

31

Monensin (MON)

35

34

23 26

22

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

Scheme 1. Structures of MON, SAL and NAR.

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MATERIALS AND METHODS

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

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

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

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

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

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

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

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Experiments confirmed that the above sample preparation method versus directly

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

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

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

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Photolysis experimental setup. Sample preparation. Reaction solutions were

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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analysis are described in SI Text S3.

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

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

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

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

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

(1)

130 131

RESULTS AND DISCUSSION

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

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

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

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

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

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only in selected cases.

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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high quantum yield compared to other veterinary pharmaceuticals.

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

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

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

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yield is demonstrated in SI Text S6 and Figure S6.

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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runoff and natural water bodies.

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

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

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

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

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

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

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

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irradiation; however, increasing the nitrate concentration resulted in diminished

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

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

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

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MON and SAL, but likely via different mechanisms.

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

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

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

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

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concentrations.

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[HO⋅]ss

∫ 2.303 ⋅ I =

λ

⋅ l ⋅ ε NO − ,λ ⋅ [NO 3− ] ⋅ Φ NO −



3

3

k Si ⋅ [Si ]

(2)

203

where

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

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

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

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l = light path length of reactor, cm

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

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

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CO32-, M

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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the transformation products (see details below).

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

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

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

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

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

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

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

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

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inhibited the photolysis of SAL.

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

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

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

248

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

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

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

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

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

253

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

254

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

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

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

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exhibited strong inhibition while SRHA and MPHA exhibited moderate enhancement on

258

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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water samples under different light irradiations, including WWTP secondary effluent

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

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

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

276

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

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

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

279

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

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

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

282

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

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

284

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

285

increased, whereas the photodegradation of SAL was not enhanced.

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

287

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

288

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

289

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

290

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

291



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



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

3

3

k HCO - /HO ⋅ ⋅ [HCO 3- ] + 3

k CO 2- /HO ⋅ ⋅ [CO 32 - ] 3

(3)

292

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

293

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

294

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

295

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

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

297

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

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

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

300

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

301 302

Product identification and phototransformation mechanisms. Products of SAL via

303

direct photolysis. Nine new significant peaks representing the major transformation

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

305

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

306

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

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

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

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

310

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

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

312

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

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

314

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

315

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

316

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

317

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

318

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

319

SAL-TP species.

320

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

321

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

322

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

323

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

324

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

325

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

326

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

327

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

328

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

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

330

poultry litter or in soil.23,62,63

331

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

332

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

333

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

334

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

335

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

336

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

337

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

338

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

339

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

340

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

341

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

342

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

343

unstable and were degraded over time.

344

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

345

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

346

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

347

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

348

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

349

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

350

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

351

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

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

353

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

354

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

355

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

356

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

357

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

358

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

359

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

360

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

361

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

362

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

363

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

364

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

365

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

366

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

367

might be generated by unknown impurities cannot be ruled out.

368

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

369

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

370

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

371

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

372

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

373

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

374

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

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

376

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

377

product yielded from other MON-TPs.

378

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

379

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

380

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

381

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

382

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

383

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

384

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

385

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

386

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

387

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

388

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

389

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

390

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

391

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

392

role in the fate of pollutants in the environment.

393 394

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

395

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

396

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

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

398

performed for SAL-TPs.

399

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

400

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

401

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

402

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

403

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

404

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

405

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

406

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

407

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

408

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

409

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

410

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

411

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

412

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

413

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

414

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

415 416

ENVIRONMENTAL IMPLICATIONS

417

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

418

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

419

Environmental conditions can greatly affect IPAs’ photodegradation. Water near

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

421

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

422

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

423

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

424

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

425

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

426

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

427

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

428

properties against target microorganisms.

429

Because photodegradation strongly depends on light irradiance and water matrices,

430

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

431

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

432

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

433

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

434

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

435

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

436

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

437

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

438

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

439

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

440

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

441

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

442

the environmental fate of IPAs.

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443 444

ASSOCIATED CONTENT

445

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

446

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

447 448

ACKNOWLEDGMENTS

449

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

450

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

451

apparatus and spectroradiometer.

452 453

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

Nitrate (mM)

pH

HCO3(mM)

CO32(mM)

WWTP secondary effluent

0.435

1.45

7.4

1.20×10-1

1.41×10-4

Simulated field runoff from PL-fertilized land

60.55

0.71

6.5

1.51×10-2

2.24×10-6

Simulated PL leachate

2074

0.01

7.2

7.59×10-2

5.62×10-5

656 657 658

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

Quantum yield (Φ)

Xenon

Sunlight

MON

SAL

MON

SAL

MON

SAL

ND

0.66±0.05

ND

0.66±0.03

ND

0.70±0.03

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

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

c

0.20±0.04 0.27±0.01

c

0.15±0.05 0.31±0.12 ND 0.08±0.02

DI + SRHA

DI + MPHA DI + PL-extractc WWTP effluentc

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

Field runoff d PL leachate e 659 660 661 662

ND

ND

ND

ND

a

(0.007)

(0.016)

ND (