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Photodegradation of Veterinary Ionophore Antibiotics under UV and Solar Irradiation Peizhe Sun, Spyros G. Pavlostathis, and Ching-Hua Huang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/es5034525 • Publication Date (Web): 24 Oct 2014 Downloaded from http://pubs.acs.org on October 28, 2014
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Environmental Science & Technology
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Photodegradation of Veterinary Ionophore Antibiotics
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under UV and Solar Irradiation
3 4
Peizhe Sun, Spyros G. Pavlostathis, Ching-Hua Huang*
5 6
School of Civil and Environmental Engineering, Georgia Institute of Technology, Atlanta,
7
Georgia 30332, United States
8 9
* Corresponding Author. Phone: 404-894-7694; Fax: 404-358-7087.
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E-mail: [email protected]
11 12
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
22
investigate the photodegradation of three commonly used IPAs, monensin (MON),
23
salinomycin (SAL) and narasin (NAR), under UV and solar irradiation. Results showed
24
that MON was persistent in deionized (DI) water matrix when exposed to UV and
25
sunlight, whereas SAL and NAR could undergo direct photolysis with a high quantum
26
yield. Water components including nitrate and dissolved organic matter had a great
27
impact on the photodegradation of IPAs. A pseudo-steady state kinetic model was
28
successfully applied to predict IPAs’ photodegradation rates in real water matrices.
29
Applying LC/MS/MS, multiple photolytic transformation products of IPAs were
30
observed and their structures proposed. The direct photolysis of SAL and NAR occurred
31
via cleavage on the ketone moiety and self-sensitized photolysis. With the presence of
32
nitrate, MON was primarily degraded by hydroxyl radicals, whereas SAL showed
33
reactivity toward both hydroxyl and nitrogen-dioxide radicals. Additionally, toxicity tests
34
showed that photodegradation of SAL eliminated its antibiotic properties against Bacillus
35
subtilis.
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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
39
2011, the annual sales of IPAs in the U.S. reached over 4.1 million kg, which made them
40
the second top-selling antimicrobial group next to tetracyclines.3 Among IPAs, monensin
41
(MON), salinomycin (SAL) and narasin (NAR) (Scheme 1) are most commonly used in
42
the livestock industry. IPAs are polyether carboxylic acids which inhibit the growth of
43
coccidia, and have shown adverse effects on Gram-positive bacteria, algae, and
44
protozoa.4-10
45
IPAs, administered in livestock feed, are mostly excreted due to poor absorption or
46
limited degradation in the animals’ digestion systems,11 and consequently are released to
47
the environment via land application of animal manure. Although IPAs are hydrophobic
48
compounds, significant amounts of IPAs can be transported with rainfall runoff from
49
manure-fertilized fields,12,13 which eventually end up in receiving waters. The residual
50
concentrations of IPAs in surface water and water treatment plants have been reported.14-
51
19
52
µg·L-1 in river water near agricultural fields.19 Watkinson et al. examined the occurrence
53
of antibiotics in different environmental water systems, among which MON and SAL
54
were detected at 94% and 21% of all collected samples (n > 84) in surface waters and
55
occasionally in wastewater treatment plant effluents.17
For example, IPAs were found at up to 0.22 µg·L-1 in surface water,16 and at 0.03−0.06
56
Photolysis is known to affect the fate of various pharmaceutical compounds in the
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aquatic environment. In surface or waste waters, degradation of IPAs may occur due to
58
exposure to solar (in natural water systems) or UV (in treatment facilities) radiation and
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such photodegradation may play an important role in affecting the overall environmental
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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
62
transformation
63
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
65
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
68
that some IPAs (e.g., SAL) may undergo direct photolysis due to the presence of the
69
carbonyl moiety in their structures.26
mechanisms
has
been
investigated
previously.
For
example,
70
To address the lack of information in the photo-behaviors of IPAs, the objectives of
71
this study were to investigate the direct and indirect photodegradation of IPAs under
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environmental conditions, measure the quantum yields, elucidate the transformation
73
products and mechanisms, and evaluate the toxicity of photo-transformation products. 5
R
HO
1
4 2
3
A O
OH
7 8
15
37
32
6
O
9
O
10
11
14 12
16
B 17
38 O
C 21
13 O 18
30
74 75 76 77
33
E
D 24
41 28 40
25 O 29
O
20 OH
OH
27
39
42
19
36
31
Monensin (MON)
35
34
23 26
22
Salinomycin (SAL) (R = -H) Narasin (NAR) (R = -CH3)
Scheme 1. Structures of MON, SAL and NAR.
78 79
MATERIALS AND METHODS
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Chemicals. Suwannee River humic acid (SRHA) was purchased from the
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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
84
prepare IPA stock solutions, individual IPA powder was weighted and pre-dissolved in
85
methanol. Then, aliquots of methanolic IPA were evaporated to dryness under vacuum
86
and re-dissolved into DI water at 5 mg·L-1, which was then stored at 4-5°C prior to use.
87
Experiments confirmed that the above sample preparation method versus directly
88
dissolving IPA in DI water yielded the same photolysis rates of IPAs, confirming that
89
residual methanol, if any, was not significant (SI Text S2 and Figure S1). Other stock
90
solution preparation and sample collection are described in SI Text S1.
91 92
Photolysis experimental setup. Sample preparation. Reaction solutions were
93
prepared with 0.5−2.5 mg·L-1 IPA (higher concentrations were used for product
94
generation) in different matrices: DI water, wastewater treatment plant (WWTP)
95
secondary effluent, simulated rainfall runoff from poultry litter (PL)-fertilized land, or
96
simulated PL leachate (matrix characteristics summarized in Table 1). The DI water
97
matrix was maintained at pH 7.0 using 5 mM phosphate buffer, whereas the other water
98
matrices were not adjusted for pH. In the experiments to investigate the effects of nitrate
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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
102
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
110
S3). Reaction solutions were kept in 10-mL quartz tubes, which were held perpendicular
111
to the light source (illustrated in SI Figure S2). For natural sunlight experiments, the
112
quartz tubes were kept vertically on the roof of a laboratory building in Atlanta, GA,
113
USA (latitude 33.8o N). Sunlight photolysis experiments were conducted in April, 2013.
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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
122
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
124
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
133
(UVC), simulated sunlight and natural sunlight irradiation. Because NAR closely
134
resembles SAL in structure with only one methyl group difference on the A ring (Scheme
135
1), NAR behaved nearly identically to SAL in acid-catalyzed hydrolysis24 and was
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expected to behave similarly to SAL in photolysis as well. Thus, NAR was investigated
137
only in selected cases.
138 139
Direct photolysis of IPAs under LP UV, simulated sunlight and natural sunlight.
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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
144
was observed, at the fastest rate under UV radiation, followed by simulated sunlight and
145
then natural sunlight (Table 2, SI Figure S4).
146
MON consists of mostly sigma bonds (C−H, C−C, C−O and O−H) except for a
147
carboxylic group at one end (Scheme 1). These structural features result in its low
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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
160
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).
174 175
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
177
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
179
due to various mechanisms, such as light-shielding effects, energy transfer to/from photo-
180
excited DOM, and/or reactions with reactive species generated by photo-sensitizers.
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Nitrate and DOM are real water components particularly considered in this study,
182
because (1) they are known to have great impact on photolysis of micropollutants in
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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
191
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
194
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
199
which estimates the steady-state HO· concentration from nitrate photolysis (Equation 2)
200
in buffered DI matrix,47,48 to predict the photolysis rates of IPAs at different nitrate
201
concentrations.
202
[HO⋅]ss
∫ 2.303 ⋅ I =
λ
⋅ l ⋅ ε NO − ,λ ⋅ [NO 3− ] ⋅ Φ NO −
∑
3
3
k Si ⋅ [Si ]
(2)
203
where
204
[HO·]ss = steady-state hydroxyl radical concentration, M
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Iλ = the irradiance received by the solution at given wavelength, Einstein·L-1·s-1
206
ελ = molar absorption coefficient of nitrate at given wavelength (M-1·cm-1)
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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
210
CO32-, M
211
kSi = second-order rate constants of co-solutes with HO·, M-1·s-1 (see SI Table S1)
212
After computing [HO·]ss, the indirect photolysis rate constants of IPAs in the
213
presence of simulated sunlight and nitrate were predicted by multiplying [HO·]ss with the
214
known second-order rate constants of IPAs with hydroxyl radicals.25 As Figure 2 shows,
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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
220
also produces several reactive nitrogen species (RNS), primarily NO2·,39,46,48 which could
221
potentially react with SAL (illustrated in SI Figure S7). Thus, the photolysis of SAL in
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the presence of nitrate was likely partly contributed by reactions with RNS. Furthermore,
223
at an increasing nitrate concentration, the RNS concentration does not necessarily
224
increase proportionally due to the possibility of radicals combining: NO· + HO· → HNO2
225
(k = 1.0×1010 M-1·s-1), NO2· + HO· → HOONO (k = 1.3×109 M-1·s-1) and NO2·+ NO2· →
226
N2O4 (k = 4.5×108 M-1·s-1).49,50 In addition, at elevated nitrate concentrations, there would
227
be more HO· which, upon reaction with IPAs in aerated solution, would produce
228
superoxide (O2-·). The latter could inhibit photonitration reactions.51 The above
229
mechanisms could explain the non-linear dependence of the rate constant on the
230
increasing nitrate concentration (Figure 2). Thus, it is hypothesized that, at a lower nitrate
231
concentration, RNS contributed to degradation of SAL, resulting in underestimation of
232
photolysis rate by the steady-state hydroxyl radical model. In contrast, at a higher nitrate
233
concentration, RNS was self-quenched yielding a diminished increase in the photolysis
234
rate of SAL. The reaction between SAL and RNS was further supported by elucidation of
235
the transformation products (see details below).
236
DOM effect. The effect of DOM on IPA photolysis was tested under simulated
237
sunlight irradiation. SRHA (3 mg C·L-1), MPHA (3 mg C·L-1), and water extracts from
238
poultry litter (PL-extract, 30 mg C·L-1) were selected because they represent commonly
239
used model DOM (i.e., SRHA and MPHA) or DOM that IPAs will likely encounter in
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agricultural fields (i.e., PL-extract). The UV-Vis spectra of the DOM working solutions
241
are shown in SI Figure S8. The degradation of IPAs was strongly influenced by DOM
242
(Table 2). Both MON and SAL degraded slightly faster in the presence of SRHA and
243
MPHA, compared with the rates in DI water matrix. However, PL-extract significantly
244
inhibited the photolysis of SAL.
245
The effects of DOM on the photolytic transformation of organic compounds have
246
been extensively studied and reported in the literature.52-54 In general, irradiation of DOM
247
may produce light-excited DOM (DOM*), singlet oxygen, and HO· species that can react
248
with organic pollutants in water and lead to enhanced pollutant degradation.54,55 On the
249
other hand, DOM may inhibit pollutants’ photolysis by light-shielding and
250
physical/chemical quenching effects, particularly at high DOM concentrations.56-58
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In this study, both enhancement and inhibition of SAL photolysis by DOM were
252
observed. All three types of DOM showed absorbance greater than 0.1 at around 285 nm
253
(SI Figure S8), suggesting that light-shielding by DOM could not be neglected in the
254
experiments. The light-shielding effect was calculated to be around 14%, 21%, and 25%
255
for SRHA, MPHA and PL-extract, respectively, under the employed experimental
256
conditions. Even though light-shielding effects by the DOM were comparable, PL-extract
257
exhibited strong inhibition while SRHA and MPHA exhibited moderate enhancement on
258
SAL photolysis. This difference was likely due to different properties of the tested DOM.
259
For example, when normalizing the light absorbance of DOM from 220 to 400 nm by
260
DOM’s DOC content (A220-400·mg C-1·L-1), PL-extract had the lowest light absorbance per
261
carbon basis. This result suggests that PL-extract contained a greater portion of non-
262
photo-reactive organic compounds which might lower PL-extract’s potential to produce
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reactive species upon irradiation and compete with IPAs to react with reactive species
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produced by photo-sensitizers. Furthermore, the inhibitory effect of PL-extract was
265
considerably greater than that attributable to light shielding, suggesting that other
266
inhibitory effects, such as physical/chemical quenching and anti-oxidation, by PL-extract
267
might also take place. Indeed, literature has shown that DOM can interact with excited
268
triplet states of ketones, leading to inhibited photolytic transformation of the ketone
269
moiety.59,60 Based on the results with PL-extract, the organic matter in runoff from PL-
270
fertilized fields is likely to exert mostly scavenging effects on IPAs’ photodegradation.
271
Photolysis in real water matrix. The photolysis of IPAs was examined in three real
272
water samples under different light irradiations, including WWTP secondary effluent
273
(under UV), simulated runoff from PL-fertilized land (under sunlight), and simulated PL
274
leachate (under sunlight). Characteristics of these water matrices are listed in Table 1.
275
Both MON and SAL photodegraded in the WWTP effluent and field runoff matrices,
276
whereas no significant loss of IPAs was detected in the PL leachate matrix for over 5
277
days. The half-lives of MON and SAL were 1.2 and 0.95 h in WWTP effluent,
278
respectively. Considering a typical UV disinfection dosage of 100 mJ·cm-2 (i.e., around 2
279
min of UV irradiation in this study), neither MON nor SAL would be efficiently removed
280
by UV treatment (less than 3% removal). Indeed, our previous study showed that, even at
281
LP UV combined with 30 mg·L-1 H2O2 AOP conditions, it required ~1000 mJ·cm-2 to
282
remove over 90% of IPAs in WWTP effluent matrix.25 In field runoff, the photolysis rate
283
constants of MON and SAL were 0.007 and 0.015 h-1 in the spring season; compared to
284
the photolysis of IPAs in DI water, the photodegradation rate of MON was significantly
285
increased, whereas the photodegradation of SAL was not enhanced.
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The overall photodegradation of IPAs in real water matrices is a result of both direct
287
and indirect photolysis. A steady-state model (Equation 3) was applied to predict IPA
288
photolysis rates with the assumptions that (1) nitrate was the main photo-reactive
289
constituent in the sample; and (2) DOM was taken into account for scavenging
290
HO· radicals. k obs = k direct + k indirect
291
∫
= Φ IPA/λ ⋅ (1 − 10 - A λ ) ⋅ f IPA, λ ⋅ I λ +
∫
k IPA/HO⋅ I λ ⋅ (1 − 10 - A λ ) ⋅ f NO − , λ ⋅ Φ NO − k IPA/HO⋅ ⋅ [IPA] + k DOM/HO ⋅ ⋅ [DOM] +
3
3
k HCO - /HO ⋅ ⋅ [HCO 3- ] + 3
k CO 2- /HO ⋅ ⋅ [CO 32 - ] 3
(3)
292
Notations were in the same manner as those in Equation (2). Additionally, Aλ was the
293
absorbance of water matrix at wavelength λ, and fIPA,λ and fNO3-,λ were the fractions of
294
light absorbed by IPAs and nitrate, respectively, at wavelength λ. Applying this model,
295
the observed rate constants of MON and SAL were estimated for different water matrices.
296
The predicted constants are close to the experimental results for the field runoff and PL
297
leachate matrices (Table 2). Although the model predicted the photodegradation rate of
298
MON in WWTP effluent well, it substantially overestimated the photodegradation rate of
299
SAL (Table 2). This overestimation might be caused by stronger physical/chemical
300
quenching of photo-excited ketone moiety of SAL by DOM in the WWTP effluent.
301 302
Product identification and phototransformation mechanisms. Products of SAL via
303
direct photolysis. Nine new significant peaks representing the major transformation
304
products of SAL (SAL-TPs) were observed in the direct photolysis experiments (SI
305
Figure S9), which had m/z (sodium adduct) of 265, 337, 447, 489, 491, 507, 519, 531,
306
and 805. Although all nine SAL-TPs were detected in the samples under UV, simulated
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sunlight and natural sunlight irradiation, the most abundant products were different for
308
each light source assuming similar MS signal response for each product (SI Figure S10).
309
Under UV irradiation, m/z 489 and 805 were the dominant SAL-TPs in the first 4 h of
310
irradiation and then were replaced by m/z 265. SAL-TPs with m/z 447, 265, 507 were
311
most abundant under simulated sunlight irradiation, whereas m/z 447 was not among the
312
major SAL-TPs in samples exposed to natural sunlight. This light-dependent generation
313
of SAL-TPs may result from two possible mechanisms. First, the absorbance of SAL
314
overlapped with UV, simulated sunlight and natural sunlight at different wavelengths (SI
315
Figure S6), which may favor specific photolytic transformation pathways to yield certain
316
dominant SAL-TPs. Indeed, dominant SAL-TPs were more similar under simulated and
317
natural sunlight, than those under LP UV. Second, the stability of SAL-TPs may be
318
different under each light source irradiation condition, which led to different dominant
319
SAL-TP species.
320
LC/MS/MS was employed to obtain more structural information of SAL-TPs. The
321
mass spectra of parent SAL and select SAL-TPs are shown in SI Figure S11. The SAL-
322
TPs with m/z of 489, 491, 507, 519 and 531 exhibited almost identical fragmentation
323
patterns below m/z 500 as that of the parent SAL. The common base peaks, m/z 431(433)
324
and 403, suggested that the products retained rings B-E, i.e., the right part of SAL
325
molecule (Scheme 1). Fragment m/z 433, instead of m/z 431, was present in some SAL-
326
TPs; this may be due to addition of an alcohol group onto C12, which rendered ring B
327
more easily to obtain a hydrogen atom from the alcohol group during fragmentation,
328
instead of eliminating a hydrogen atom from C14.61 It is also noteworthy that the
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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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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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(62)
632
salinomycin. J. Antibiot. 1987, 40 (3), 388-390.
633
(63)
634
transformation of anticoccidials in soil using a lab-scale bio-reactor as a precursor-tool.
635
Chemosphere 2012, 86 (2), 212-215.
636
(64)
637
of organic molecules. Photochem. Photobiol. 2012, 88 (4), 1033-1033.
638
(65)
639
Ajassa, R.; Olariu, R. I.; Arsene, C. Sources and sinks of hydroxyl radicals upon
640
irradiation of natural water samples. Environ. Sci. Technol. 2006, 40 (12), 3775-3781.
641
(66)
642
New processes in the environmental chemistry of nitrite. 2. The role of hydrogen
643
peroxide. Environ. Sci. Technol. 2003, 37 (20), 4635-4641.
Loutfy, R. O.; Yip, R. W. Triplet-state of ketones in solution - role of charge-
Volmer, D. A.; Lock, C. M. Electrospray ionization and collision-induced
Vertesy, L.; Heil, K.; Fehlhaber, H. W.; Ziegler, W. Microbial decomposition of
Hansen, M.; Bjorklund, E.; Krogh, K. A.; Brandt, A.; Halling-Sorensen, B. Biotic
Turro, N. J.; Ramamurthy, V.; Scaiano, J. C. Modern molecular photochemistry
Vione, D.; Falletti, G.; Maurino, V.; Minero, C.; Pelizzetti, E.; Malandrino, M.;
Vione, D.; Maurino, V.; Minero, C.; Borghesi, D.; Lucchiari, M.; Pelizzetti, E.
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(67)
Vione, D.; Maurino, V.; Minero, C.; Pelizzetti, E. New processes in the
645
environmental chemistry of nitrite: nitration of phenol upon nitrite photoinduced
646
oxidation. Environ. Sci. Technol. 2002, 36 (4), 669-676.
647
(68)
648
nitrophenols upon UV irradiation of phenol and nitrate in aqueous solutions and in TiO2
649
aqueous suspensions. Chemosphere 2001, 44 (2), 237-248.
650
(69)
651
Dekker: New York, 1982.
652
(70)
653
control of coccidiosis in poultry. Poultry Sci. 2010, 89 (9), 1788-1801.
Vione, D.; Maurino, V.; Minero, C.; Vincenti, M.; Pelizzetti, E. Formation of
Westley, J. W. Polyether Antibiotics: Naturally Occurring Acid Ionophores; M.
Chapman, H. D.; Jeffers, T. K.; Williams, R. B. Forty years of monensin for the
654
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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 (
Article
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
2
under UV and Solar Irradiation
3 4
Peizhe Sun, Spyros G. Pavlostathis, Ching-Hua Huang*
5 6
School of Civil and Environmental Engineering, Georgia Institute of Technology, Atlanta,
7
Georgia 30332, United States
8 9
* Corresponding Author. Phone: 404-894-7694; Fax: 404-358-7087.
10
E-mail: [email protected]
11 12
Revised manuscript submitted to
13
Environmental Science & Technology
14 15
October 24, 2014
16
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ABSTRACT
18
The veterinary ionophore antibiotics (IPAs) are extensively used as coccidiostats and
19
growth promoters, and are released to the environment via land application of animal
20
waste. Due to their propensity to be transported with runoff, IPAs likely end up in surface
21
waters where they are subject to photodegradation. This study is among the first to
22
investigate the photodegradation of three commonly used IPAs, monensin (MON),
23
salinomycin (SAL) and narasin (NAR), under UV and solar irradiation. Results showed
24
that MON was persistent in deionized (DI) water matrix when exposed to UV and
25
sunlight, whereas SAL and NAR could undergo direct photolysis with a high quantum
26
yield. Water components including nitrate and dissolved organic matter had a great
27
impact on the photodegradation of IPAs. A pseudo-steady state kinetic model was
28
successfully applied to predict IPAs’ photodegradation rates in real water matrices.
29
Applying LC/MS/MS, multiple photolytic transformation products of IPAs were
30
observed and their structures proposed. The direct photolysis of SAL and NAR occurred
31
via cleavage on the ketone moiety and self-sensitized photolysis. With the presence of
32
nitrate, MON was primarily degraded by hydroxyl radicals, whereas SAL showed
33
reactivity toward both hydroxyl and nitrogen-dioxide radicals. Additionally, toxicity tests
34
showed that photodegradation of SAL eliminated its antibiotic properties against Bacillus
35
subtilis.
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INTRODUCTION
37
Ionophore antibiotics (IPAs) are among the most common veterinary pharmaceuticals
38
used to prevent coccidiosis and promote digestion efficiency for meat production.1,2 In
39
2011, the annual sales of IPAs in the U.S. reached over 4.1 million kg, which made them
40
the second top-selling antimicrobial group next to tetracyclines.3 Among IPAs, monensin
41
(MON), salinomycin (SAL) and narasin (NAR) (Scheme 1) are most commonly used in
42
the livestock industry. IPAs are polyether carboxylic acids which inhibit the growth of
43
coccidia, and have shown adverse effects on Gram-positive bacteria, algae, and
44
protozoa.4-10
45
IPAs, administered in livestock feed, are mostly excreted due to poor absorption or
46
limited degradation in the animals’ digestion systems,11 and consequently are released to
47
the environment via land application of animal manure. Although IPAs are hydrophobic
48
compounds, significant amounts of IPAs can be transported with rainfall runoff from
49
manure-fertilized fields,12,13 which eventually end up in receiving waters. The residual
50
concentrations of IPAs in surface water and water treatment plants have been reported.14-
51
19
52
µg·L-1 in river water near agricultural fields.19 Watkinson et al. examined the occurrence
53
of antibiotics in different environmental water systems, among which MON and SAL
54
were detected at 94% and 21% of all collected samples (n > 84) in surface waters and
55
occasionally in wastewater treatment plant effluents.17
For example, IPAs were found at up to 0.22 µg·L-1 in surface water,16 and at 0.03−0.06
56
Photolysis is known to affect the fate of various pharmaceutical compounds in the
57
aquatic environment. In surface or waste waters, degradation of IPAs may occur due to
58
exposure to solar (in natural water systems) or UV (in treatment facilities) radiation and
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such photodegradation may play an important role in affecting the overall environmental
60
fate of IPAs. To date, however, few studies have evaluated the susceptibility of IPAs to
61
photo-induced degradation. In comparison, the degradation of IPAs via other
62
transformation
63
biodegradation of IPAs was reported in several studies with half-lives of 3−5 days in
64
soil.20-23 IPA’s instability to acid-catalyzed hydrolysis was investigated by our recent
65
study which elucidated transformation mechanisms and products.24 The degradation of
66
IPAs by UV/H2O2 advanced oxidation process (AOP) was also examined by us
67
recently.25 Although studies on the photodegradation of IPAs are scarce, it is expected
68
that some IPAs (e.g., SAL) may undergo direct photolysis due to the presence of the
69
carbonyl moiety in their structures.26
mechanisms
has
been
investigated
previously.
For
example,
70
To address the lack of information in the photo-behaviors of IPAs, the objectives of
71
this study were to investigate the direct and indirect photodegradation of IPAs under
72
environmental conditions, measure the quantum yields, elucidate the transformation
73
products and mechanisms, and evaluate the toxicity of photo-transformation products. 5
R
HO
1
4 2
3
A O
OH
7 8
15
37
32
6
O
9
O
10
11
14 12
16
B 17
38 O
C 21
13 O 18
30
74 75 76 77
33
E
D 24
41 28 40
25 O 29
O
20 OH
OH
27
39
42
19
36
31
Monensin (MON)
35
34
23 26
22
Salinomycin (SAL) (R = -H) Narasin (NAR) (R = -CH3)
Scheme 1. Structures of MON, SAL and NAR.
78 79
MATERIALS AND METHODS
80
Chemicals. Suwannee River humic acid (SRHA) was purchased from the
81
International Humic Substances Society (IHSS, St. Paul, MN, USA). Another type of
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humic acid (MPHA) was purchased from MP Biomedicals (Solon, OH, USA). Sources of
83
other chemicals and reagents are provided in the Supporting Information (SI) Text S1. To
84
prepare IPA stock solutions, individual IPA powder was weighted and pre-dissolved in
85
methanol. Then, aliquots of methanolic IPA were evaporated to dryness under vacuum
86
and re-dissolved into DI water at 5 mg·L-1, which was then stored at 4-5°C prior to use.
87
Experiments confirmed that the above sample preparation method versus directly
88
dissolving IPA in DI water yielded the same photolysis rates of IPAs, confirming that
89
residual methanol, if any, was not significant (SI Text S2 and Figure S1). Other stock
90
solution preparation and sample collection are described in SI Text S1.
91 92
Photolysis experimental setup. Sample preparation. Reaction solutions were
93
prepared with 0.5−2.5 mg·L-1 IPA (higher concentrations were used for product
94
generation) in different matrices: DI water, wastewater treatment plant (WWTP)
95
secondary effluent, simulated rainfall runoff from poultry litter (PL)-fertilized land, or
96
simulated PL leachate (matrix characteristics summarized in Table 1). The DI water
97
matrix was maintained at pH 7.0 using 5 mM phosphate buffer, whereas the other water
98
matrices were not adjusted for pH. In the experiments to investigate the effects of nitrate
99
and dissolved organic matter (DOM), aliquots of nitrate or DOM stock solutions were
100
spiked into the DI water matrix to achieve the target concentrations.
101
UV Photolysis. Photolysis of IPAs with UV radiation was studied using a similar
102
setup as previously described,25 and shown in SI Figure S2. Experiments were conducted
103
in a magnetically stirred 100-mL cylindrical quartz reactor kept in a photo-chamber
104
equipped with a 4-W low pressure (LP) UV lamp (G4T5 Hg lamp, Philips TUV4W)
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peaking at 254 nm at ambient temperature (22oC). Reaction was initiated by exposing the
106
solution to UV irradiation. A sample aliquot was taken at each time interval and stored in
107
a 2-mL amber glass vial prior to LC/MS analysis.
108
Simulated and natural sunlight photolysis. Simulated sunlight was generated by a
109
300-W Xenon lamp (PerkinElmer, PE300BF, full spectral output is shown in SI Figure
110
S3). Reaction solutions were kept in 10-mL quartz tubes, which were held perpendicular
111
to the light source (illustrated in SI Figure S2). For natural sunlight experiments, the
112
quartz tubes were kept vertically on the roof of a laboratory building in Atlanta, GA,
113
USA (latitude 33.8o N). Sunlight photolysis experiments were conducted in April, 2013.
114
Chemical and instrumental analysis. IPAs were analyzed by an Agilent 1100 Series
115
high performance liquid chromatography mass spectrometry (HPLC/MS) system (Agilent,
116
Palo Alto, CA) with a reversed-phase Ascentis RP-amide column (2.1 × 150 mm, 3 µm).
117
The mobile phase contained (A) 0.1% formic acid containing 25% acetonitrile, and (B) a
118
mixture of 50/50 acetonitrile/methanol (v/v). The analytical methods were described in
119
detail previously.27 For elucidation of photo-transformation products, a LC/MS/MS unit
120
(Agilent 1260 Infinity LC system, 6410 Triple Quad MSD, Agilent, Palo Alto, CA) was
121
employed. The LC conditions were similar to those of the HPLC/MS system described
122
above. The MS parameters were set up with fragmentation voltage of 135 V, and
123
collision-induced-dissociation (CID) energy of 50 eV. Other details on instrumental
124
analysis are described in SI Text S3.
125
Toxicity tests for transformation products. The growth of Bacillus subtilis (ATCC
126
6633) was shown to be sensitive to IPA concentration and thus this microorganism was
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chosen for the toxicity tests.10,24 The growth inhibition is expressed in Equation 1, in
128
which OD is the optical density. Details of toxicity tests are described in SI Text S4.
129
OD 600 (Sample) × 100 Growth Inhibition (%) = 1 − OD 600 (Control)
(1)
130 131
RESULTS AND DISCUSSION
132
The photodegradation of MON and SAL was investigated extensively under LP UV
133
(UVC), simulated sunlight and natural sunlight irradiation. Because NAR closely
134
resembles SAL in structure with only one methyl group difference on the A ring (Scheme
135
1), NAR behaved nearly identically to SAL in acid-catalyzed hydrolysis24 and was
136
expected to behave similarly to SAL in photolysis as well. Thus, NAR was investigated
137
only in selected cases.
138 139
Direct photolysis of IPAs under LP UV, simulated sunlight and natural sunlight.
140
Photolysis of IPAs due to direct absorption of light was investigated in DI water buffered
141
by 5 mM phosphate at pH 7. In all dark controls, no significant loss of IPAs was
142
observed for over one week (data not shown). MON was resistant to photodegradation in
143
DI water matrix under exposure to any light source. In contrast, the degradation of SAL
144
was observed, at the fastest rate under UV radiation, followed by simulated sunlight and
145
then natural sunlight (Table 2, SI Figure S4).
146
MON consists of mostly sigma bonds (C−H, C−C, C−O and O−H) except for a
147
carboxylic group at one end (Scheme 1). These structural features result in its low
148
absorption of light above 220 nm (Figure 1). However, the additional C=C bond
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(C18−C19) and carbonyl group (C11−O) of SAL (and NAR) contribute to two significant
150
absorbance bands, which peak at below 230 nm and 285 nm, respectively (Figure 1). The
151
first-order rate constant of photolysis of NAR under UV irradiation was 0.47±0.01 h-1,
152
nearly identical to that of SAL (0.46 h-1, Table 2) due to their similar molecular structures.
153
The half-life of SAL in DI water with sunlight radiation (in April) was 53.3±4.1 h,
154
which is relatively long compared to other commonly used veterinary pharmaceuticals.
155
For example, the half-life of tetracycline by photolysis was around 44 min in July;28 and
156
that of sulfamethoxazole was around 10.4 h in summer.29 However, considering the fairly
157
low molar extinction coefficient of SAL (Figure 1), it suggests that SAL has a relatively
158
high quantum yield compared to other veterinary pharmaceuticals.
159
Quantum yield of SAL. The quantum yields of SAL under different light radiation are
160
shown in Table 2. The determination of photo-parameters with respect to the
161
experimental setup is described in SI Text S5 and Figure S5. The calculation of quantum
162
yield is demonstrated in SI Text S6 and Figure S6.
163
The quantum yield of SAL was similar under the three different light sources.
164
Comparing to other pharmaceuticals, the quantum yield of SAL is considerably higher.
165
For example, the quantum yields of tetracycline (0.00024−0.002),28 naproxen (0.012),30
166
diclofenac (0.031−0.22),31,32 carbamazepine (0.00013),33 levofloxacin (0.00008),34
167
amoxicillin (0.0045-0.0060),35 and sulfamethoxazole (0.02)33 are all in the range of
168
0.001−0.22 under simulated sunlight or natural sunlight, which are much lower than that
169
of SAL. On the other hand, some ketone-containing compounds, such as penton-2-one,
170
hexan-2-one, 2,2-dimethylheptan-3-one, 3,3-dimethylbutan-2-one, 2,2-dimethylhexan-3-
171
one, and 2,2,4,4-tetramethylpentan-3-one have quantum yield values of 0.27−0.71.36,37
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172
The similar range of quantum yields of SAL and the ketone-containing compounds
173
suggests that the photo-reactive part of SAL is likely the ketone moiety (C11−O).
174 175
Indirect photolysis of IPAs. Although MON is stable under all three light sources in
176
DI matrix, indirect photolysis of MON is expected in real water matrices via reactions
177
with reactive species generated by natural photo-sensitizers. For SAL, the observed
178
photolysis rate in real water matrix may also be influenced by different water components
179
due to various mechanisms, such as light-shielding effects, energy transfer to/from photo-
180
excited DOM, and/or reactions with reactive species generated by photo-sensitizers.
181
Nitrate and DOM are real water components particularly considered in this study,
182
because (1) they are known to have great impact on photolysis of micropollutants in
183
water systems;38-41 and (2) they are major components in WWTP effluent, agricultural
184
runoff and natural water bodies.
185
Nitrate effect. Nitrate concentrations in WWTP effluent and agricultural runoff could
186
reach 10 mg·L-1 as N or higher. In this study, 0−10 mg·L-1 nitrate as sodium nitrate was
187
spiked into DI water matrix under simulated sunlight exposure to investigate the
188
influence of nitrate on IPA photolysis. As Figure 2 shows, the presence of nitrate
189
enhanced the photodegradation of both MON and SAL. The first-order degradation rate
190
constant of MON increased linearly with increasing nitrate concentration (R2 > 0.99). The
191
presence of nitrate also enhanced the degradation of SAL under simulated sunlight
192
irradiation; however, increasing the nitrate concentration resulted in diminished
193
enhancement of the degradation rate of SAL. Because the presence of nitrate did not raise
194
the light absorbance to greater than 0.1, light-shielding effect was negligible.42 The
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195
combined results suggest that the presence of nitrate enhanced the photolysis of both
196
MON and SAL, but likely via different mechanisms.
197
Nitrate is well known to produce hydroxyl radicals under either UV or sunlight
198
exposure.43-46 Therefore, attempts were made to apply an established numerical model
199
which estimates the steady-state HO· concentration from nitrate photolysis (Equation 2)
200
in buffered DI matrix,47,48 to predict the photolysis rates of IPAs at different nitrate
201
concentrations.
202
[HO⋅]ss
∫ 2.303 ⋅ I =
λ
⋅ l ⋅ ε NO − ,λ ⋅ [NO 3− ] ⋅ Φ NO −
∑
3
3
k Si ⋅ [Si ]
(2)
203
where
204
[HO·]ss = steady-state hydroxyl radical concentration, M
205
Iλ = the irradiance received by the solution at given wavelength, Einstein·L-1·s-1
206
ελ = molar absorption coefficient of nitrate at given wavelength (M-1·cm-1)
207
l = light path length of reactor, cm
208
Φ = quantum yield of HO· production from nitrate (0.017 M·Einstein-1)
209
[Si] = concentrations of co-solutes (Si), including IPA, H2PO4-, HPO42-, HCO3- and
210
CO32-, M
211
kSi = second-order rate constants of co-solutes with HO·, M-1·s-1 (see SI Table S1)
212
After computing [HO·]ss, the indirect photolysis rate constants of IPAs in the
213
presence of simulated sunlight and nitrate were predicted by multiplying [HO·]ss with the
214
known second-order rate constants of IPAs with hydroxyl radicals.25 As Figure 2 shows,
215
the predicted rate constants agreed well with the experimental data, which suggests that
216
indirect photolysis of MON with nitrate was primarily via the reaction with HO·.
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217
However, the model underestimated the degradation rates of SAL at 0.2−1 mg·L-1
218
NaNO3 (Figure 2), which suggests reactive species other than HO· may play a part in the
219
indirect photolysis of SAL. Indeed, besides generation of HO·, the photolysis of nitrate
220
also produces several reactive nitrogen species (RNS), primarily NO2·,39,46,48 which could
221
potentially react with SAL (illustrated in SI Figure S7). Thus, the photolysis of SAL in
222
the presence of nitrate was likely partly contributed by reactions with RNS. Furthermore,
223
at an increasing nitrate concentration, the RNS concentration does not necessarily
224
increase proportionally due to the possibility of radicals combining: NO· + HO· → HNO2
225
(k = 1.0×1010 M-1·s-1), NO2· + HO· → HOONO (k = 1.3×109 M-1·s-1) and NO2·+ NO2· →
226
N2O4 (k = 4.5×108 M-1·s-1).49,50 In addition, at elevated nitrate concentrations, there would
227
be more HO· which, upon reaction with IPAs in aerated solution, would produce
228
superoxide (O2-·). The latter could inhibit photonitration reactions.51 The above
229
mechanisms could explain the non-linear dependence of the rate constant on the
230
increasing nitrate concentration (Figure 2). Thus, it is hypothesized that, at a lower nitrate
231
concentration, RNS contributed to degradation of SAL, resulting in underestimation of
232
photolysis rate by the steady-state hydroxyl radical model. In contrast, at a higher nitrate
233
concentration, RNS was self-quenched yielding a diminished increase in the photolysis
234
rate of SAL. The reaction between SAL and RNS was further supported by elucidation of
235
the transformation products (see details below).
236
DOM effect. The effect of DOM on IPA photolysis was tested under simulated
237
sunlight irradiation. SRHA (3 mg C·L-1), MPHA (3 mg C·L-1), and water extracts from
238
poultry litter (PL-extract, 30 mg C·L-1) were selected because they represent commonly
239
used model DOM (i.e., SRHA and MPHA) or DOM that IPAs will likely encounter in
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240
agricultural fields (i.e., PL-extract). The UV-Vis spectra of the DOM working solutions
241
are shown in SI Figure S8. The degradation of IPAs was strongly influenced by DOM
242
(Table 2). Both MON and SAL degraded slightly faster in the presence of SRHA and
243
MPHA, compared with the rates in DI water matrix. However, PL-extract significantly
244
inhibited the photolysis of SAL.
245
The effects of DOM on the photolytic transformation of organic compounds have
246
been extensively studied and reported in the literature.52-54 In general, irradiation of DOM
247
may produce light-excited DOM (DOM*), singlet oxygen, and HO· species that can react
248
with organic pollutants in water and lead to enhanced pollutant degradation.54,55 On the
249
other hand, DOM may inhibit pollutants’ photolysis by light-shielding and
250
physical/chemical quenching effects, particularly at high DOM concentrations.56-58
251
In this study, both enhancement and inhibition of SAL photolysis by DOM were
252
observed. All three types of DOM showed absorbance greater than 0.1 at around 285 nm
253
(SI Figure S8), suggesting that light-shielding by DOM could not be neglected in the
254
experiments. The light-shielding effect was calculated to be around 14%, 21%, and 25%
255
for SRHA, MPHA and PL-extract, respectively, under the employed experimental
256
conditions. Even though light-shielding effects by the DOM were comparable, PL-extract
257
exhibited strong inhibition while SRHA and MPHA exhibited moderate enhancement on
258
SAL photolysis. This difference was likely due to different properties of the tested DOM.
259
For example, when normalizing the light absorbance of DOM from 220 to 400 nm by
260
DOM’s DOC content (A220-400·mg C-1·L-1), PL-extract had the lowest light absorbance per
261
carbon basis. This result suggests that PL-extract contained a greater portion of non-
262
photo-reactive organic compounds which might lower PL-extract’s potential to produce
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reactive species upon irradiation and compete with IPAs to react with reactive species
264
produced by photo-sensitizers. Furthermore, the inhibitory effect of PL-extract was
265
considerably greater than that attributable to light shielding, suggesting that other
266
inhibitory effects, such as physical/chemical quenching and anti-oxidation, by PL-extract
267
might also take place. Indeed, literature has shown that DOM can interact with excited
268
triplet states of ketones, leading to inhibited photolytic transformation of the ketone
269
moiety.59,60 Based on the results with PL-extract, the organic matter in runoff from PL-
270
fertilized fields is likely to exert mostly scavenging effects on IPAs’ photodegradation.
271
Photolysis in real water matrix. The photolysis of IPAs was examined in three real
272
water samples under different light irradiations, including WWTP secondary effluent
273
(under UV), simulated runoff from PL-fertilized land (under sunlight), and simulated PL
274
leachate (under sunlight). Characteristics of these water matrices are listed in Table 1.
275
Both MON and SAL photodegraded in the WWTP effluent and field runoff matrices,
276
whereas no significant loss of IPAs was detected in the PL leachate matrix for over 5
277
days. The half-lives of MON and SAL were 1.2 and 0.95 h in WWTP effluent,
278
respectively. Considering a typical UV disinfection dosage of 100 mJ·cm-2 (i.e., around 2
279
min of UV irradiation in this study), neither MON nor SAL would be efficiently removed
280
by UV treatment (less than 3% removal). Indeed, our previous study showed that, even at
281
LP UV combined with 30 mg·L-1 H2O2 AOP conditions, it required ~1000 mJ·cm-2 to
282
remove over 90% of IPAs in WWTP effluent matrix.25 In field runoff, the photolysis rate
283
constants of MON and SAL were 0.007 and 0.015 h-1 in the spring season; compared to
284
the photolysis of IPAs in DI water, the photodegradation rate of MON was significantly
285
increased, whereas the photodegradation of SAL was not enhanced.
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286
The overall photodegradation of IPAs in real water matrices is a result of both direct
287
and indirect photolysis. A steady-state model (Equation 3) was applied to predict IPA
288
photolysis rates with the assumptions that (1) nitrate was the main photo-reactive
289
constituent in the sample; and (2) DOM was taken into account for scavenging
290
HO· radicals. k obs = k direct + k indirect
291
∫
= Φ IPA/λ ⋅ (1 − 10 - A λ ) ⋅ f IPA, λ ⋅ I λ +
∫
k IPA/HO⋅ I λ ⋅ (1 − 10 - A λ ) ⋅ f NO − , λ ⋅ Φ NO − k IPA/HO⋅ ⋅ [IPA] + k DOM/HO ⋅ ⋅ [DOM] +
3
3
k HCO - /HO ⋅ ⋅ [HCO 3- ] + 3
k CO 2- /HO ⋅ ⋅ [CO 32 - ] 3
(3)
292
Notations were in the same manner as those in Equation (2). Additionally, Aλ was the
293
absorbance of water matrix at wavelength λ, and fIPA,λ and fNO3-,λ were the fractions of
294
light absorbed by IPAs and nitrate, respectively, at wavelength λ. Applying this model,
295
the observed rate constants of MON and SAL were estimated for different water matrices.
296
The predicted constants are close to the experimental results for the field runoff and PL
297
leachate matrices (Table 2). Although the model predicted the photodegradation rate of
298
MON in WWTP effluent well, it substantially overestimated the photodegradation rate of
299
SAL (Table 2). This overestimation might be caused by stronger physical/chemical
300
quenching of photo-excited ketone moiety of SAL by DOM in the WWTP effluent.
301 302
Product identification and phototransformation mechanisms. Products of SAL via
303
direct photolysis. Nine new significant peaks representing the major transformation
304
products of SAL (SAL-TPs) were observed in the direct photolysis experiments (SI
305
Figure S9), which had m/z (sodium adduct) of 265, 337, 447, 489, 491, 507, 519, 531,
306
and 805. Although all nine SAL-TPs were detected in the samples under UV, simulated
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sunlight and natural sunlight irradiation, the most abundant products were different for
308
each light source assuming similar MS signal response for each product (SI Figure S10).
309
Under UV irradiation, m/z 489 and 805 were the dominant SAL-TPs in the first 4 h of
310
irradiation and then were replaced by m/z 265. SAL-TPs with m/z 447, 265, 507 were
311
most abundant under simulated sunlight irradiation, whereas m/z 447 was not among the
312
major SAL-TPs in samples exposed to natural sunlight. This light-dependent generation
313
of SAL-TPs may result from two possible mechanisms. First, the absorbance of SAL
314
overlapped with UV, simulated sunlight and natural sunlight at different wavelengths (SI
315
Figure S6), which may favor specific photolytic transformation pathways to yield certain
316
dominant SAL-TPs. Indeed, dominant SAL-TPs were more similar under simulated and
317
natural sunlight, than those under LP UV. Second, the stability of SAL-TPs may be
318
different under each light source irradiation condition, which led to different dominant
319
SAL-TP species.
320
LC/MS/MS was employed to obtain more structural information of SAL-TPs. The
321
mass spectra of parent SAL and select SAL-TPs are shown in SI Figure S11. The SAL-
322
TPs with m/z of 489, 491, 507, 519 and 531 exhibited almost identical fragmentation
323
patterns below m/z 500 as that of the parent SAL. The common base peaks, m/z 431(433)
324
and 403, suggested that the products retained rings B-E, i.e., the right part of SAL
325
molecule (Scheme 1). Fragment m/z 433, instead of m/z 431, was present in some SAL-
326
TPs; this may be due to addition of an alcohol group onto C12, which rendered ring B
327
more easily to obtain a hydrogen atom from the alcohol group during fragmentation,
328
instead of eliminating a hydrogen atom from C14.61 It is also noteworthy that the
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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 (
