Environ. Sci. Technol. 2010, 44, 2400–2405
Aquatic Photochemistry of Fluoroquinolone Antibiotics: Kinetics, Pathways, and Multivariate Effects of Main Water Constituents LINKE GE, JINGWEN CHEN,* XIAOXUAN WEI, SIYU ZHANG, XIANLIANG QIAO, XIYUN CAI, AND QING XIE Key Laboratory of Industrial Ecology and Environmental Engineering (MOE), Department of Environmental Science and Technology, Dalian University of Technology, Linggong Road 2, Dalian 116024, P. R. China
Received September 21, 2009. Revised manuscript received December 29, 2009. Accepted February 8, 2010.
The ubiquity of fluoroquinolone antibiotics (FQs) in surface waters urges insights into their fate in the aqueous euphotic zone. In this study, eight FQs (ciprofloxacin, danofloxacin, levofloxacin, sarafloxacin, difloxacin, enrofloxacin, gatifloxacin, and balofloxacin) were exposed to simulated sunlight, and their photodegradation was observed to follow apparent firstorder kinetics. Based on the determined photolytic quantum yields, solar photodegradation half-lives for the FQs in pure water and at 45° N latitude were calculated to range from 1.25 min for enrofloxacin to 58.0 min for balofloxacin, suggesting that FQs would intrinsically photodegrade fast in sunlit surface waters. However, we found freshwater and seawater constituents inhibited their photodegradation. The inhibition was further explored by a central composite design using sarafloxacin and gatifloxacin as representatives. Humic acids (HA), Fe(III), NO3-, and HA-Cl- interaction inhibited the photodegradation, as they mainly acted as radiation filters and/or scavengers for reactive oxygen species. The photodegradation product identification and ROS scavenging experiments indicated that the FQs underwent both direct photolysis and self-sensitized photo-oxidation via •OH and 1O2. Piperazinyl N4-dealkylation was primary for N4-alkylated FQs, whereas decarboxylation and defluorination were comparatively important for the other FQs. These results are of importance toward the goal of assessing the persistence of FQs in surface waters.
Introduction Pharmaceuticals, especially antibiotics, have received increasing attention as aqueous micropollutants with their environmental fate and impact to be understood (1). Fluoroquinolones (FQs) are a large class of antibiotics that are widely used in aquiculture, livestock husbandry, and human prescription (2, 3). Until recently, the ubiquity of FQs in surface waters has been found and confirmed by studies conducted in China (4-6), Europe (7-9), and the U.S. (10). To assess the fate and risk of FQs in aquatic systems, quantification of the pertinent processes that determine their fate and effect is a prerequisite. * Corresponding author phone/fax: +86-411-84706269; e-mail:
[email protected]. 2400
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The environmental fate of FQs is influenced by three primary mechanisms, photodegradation, adsorption, and biodegradation (11, 12), with photodegradation and adsorption to be most significant in aquatic systems (12, 13). Knapp et al. (12) investigated the fate of enrofloxacin (ENR) in outdoor mesocosms under different light conditions, and found that ENR was transformed into ciprofloxacin (CIP). They further examined CIP disappearance rate in surface waters using both laboratory and field systems under different light, and dissolved (DOC) and particulate organic carbon (POC) conditions to determine when photodegradation versus adsorption dominates CIP fate, and found CIP rapidly photodegraded (t1/2 ≈ 1.5 h) when POC levels were low (13). Araki et al. (14) investigated the photochemical behavior of sitafloxacin in aqueous solutions and found that matrix pH, Cl-, and Br- had an impact on the photodegradation kinetics and product distribution. There are many variables affecting the photofate of pollutants in aquatic systems, such as multiplicate water constituents and different light conditions (12, 15, 16). Besides DOC, POC, Cl-, Br-, pH and light conditions, Fe(III) and NO3- may also have an impact on the photodegradation kinetics of FQs due to their known photoreactivity in sunlit surface waters (17, 18). As these photoreactive species coexist in natural waters, it is necessary to consider the multivariate effects of the water constituents (17, 19). Organic pollutants can undergo not only direct or indirect photodegradation (20-22) but also self-sensitized photooxidation (23-26). Clarifying the photolytic pathways and mechanisms is of importance to understand the risks they may pose (27). Albini and Monti (28) suggested that FQs underwent defluorination, decarboxylation, and side-chain degradation in aqueous solutions. These pathways strongly depend on the matrix conditions and light sources (12, 29, 30). In addition, some previous studies indicated the generation of reactive oxygen species (ROS, such as •OH and 1O2) in light-irradiated FQ solutions (31-33). However, it is not clear whether FQs underwent self-sensitized photo-oxidation via ROS (28). This study investigates the photochemical behavior of eight FQs, including CIP, danofloxacin (DAN), levofloxacin (LEV), sarafloxacin (SAR), difloxacin (DIF), ENR, gatifloxacin (GAT), and balofloxacin (BAL). Photodegradation research under environmentally relevant conditions for most of the FQs (e.g., DAN, SAR, GAT, and BAL) is deficient in the previous studies. SAR and GAT were selected as representatives to assess the multivariate effects of water ingredients [Fe(III), NO3-, HA and Cl-] on the photolysis. Moreover, intermediates and degradation products were identified to explain the main pathways, and the toxicity evolvement during the phototransformation was assessed.
Experimental Section Chemicals. The eight FQs were obtained from different suppliers with >98% purity (Supporting Information (SI) Table S1). Humic acids (HA, Fluka no. 53680) were purchased from Sigma Aldrich, Inc. (Milwaukee, WI). All the organic solvents (HPLC grade) were purchased from Tedia Inc. Other reagents were of guaranteed grade and used as received. Ultra pure water was obtained with a Millipore-Milli Q system. Freshwater and seawater were collected from a local protected reservoir and the Yellow Sea, respectively, filtered through 0.22 µm filters, and stored at -20 °C until use. Local humic acids (L-HA) were extracted from the freshwater of the local reservoir following the method recommended by the Inter10.1021/es902852v
2010 American Chemical Society
Published on Web 03/05/2010
TABLE 2. Photolytic Quantum Yields (Φ) and Solar Photodegradation Half-Lives (t1/2) for the FQs in Pure Water and at 45° N Latitude t1/2 (min)b no.
FIGURE 1. Ultraviolet absorption spectra of 5 µM sarafloxacin in three waters and irradiance spectrum of the simulated sunlight (relative intensity).
TABLE 1. Factors, Levels and Initial Concentrations in the Four-Factor Central Composite Design factor (units)
1 2 3 4 5 6 7 8
Φa
FQs CIP DAN LEV SAR DIF ENR GAT BAL
-2
(5.48 ( 1.92) × 10 (3.03 ( 0.54) × 10-2 (8.26 ( 1.08) × 10-3 (3.97 ( 1.10) × 10-2 (3.13 ( 0.41) × 10-2 (6.97 ( 1.41) × 10-2 (5.94 ( 0.95) × 10-3 (4.72 ( 0.56) × 10-3
summer
winter
1.80 1.76 8.94 2.34 2.22 1.25 11.0 12.5
8.82 7.84 42.2 11.5 10.5 6.11 51.8 58.0
a Mean ( 95% confidence interval, n ) 3. b t1/2 were calculated using tabular solar intensities at noon of summer and winter, assuming continuous irradiation. The intensity value at every specific wavelength at 45° latitude was taken as the average of the 40° and 50° latitude values.
factor concentrations
coded factor levels -2 -1 0 1 2 Fe(III) (µM) 0.00 1.00 2.00 3.00 4.00 NO3- (µM) 0.00 10.0 20.0 30.0 40.0 HA (mg C/L) 0.00 2.50 5.00 7.50 10.0 Cl- (M) 0.00 0.125 0.250 0.375 0.500
national Humic Substances Society (http://www.ihss. gatech.edu/). Photodegradation Experiments. An XPA-1 merry-goround photochemical reactor (Xujiang Electromechanical Plant, Nanjing, China) with a pyrex-well cooled and filtered xenon lamp (1000 W) to simulate sunlight was used. To facilitate observations of photolytic curves and accumulation of degradation intermediates for the chemical analysis, only one lamp was used so that the FQs could photolyze neither too fast nor too slow. The irradiance spectrum of the light source (Figure 1 and SI Figure S1) was measured with an Acton SP-300 monochromator. The light intensity (290-420 nm) at the reaction solutions was 0.83 mW/cm2. Pure water, freshwater and seawater were used for the experiments, with the initial concentration (C0) of each FQ being lower than 5 µM. Quantum yields (Φ) in pure water were measured using p-nitroanisole/pyridine as a chemical actinometer (34, 35). To test whether the FQs underwent the self-sensitized photodegradation via ROS, scavenging experiments were performed using isopropanol and sodium azide (NaN3). Isopropanol is the quencher of •OH (36), and NaN3 is the quencher of 1O2 and •OH (36, 37). They were frequently used in photochemical studies to investigate the reaction mechanisms (38-40). A four-factor central composite design was used to investigate the multivariate effects of Fe(III), HA, NO3-, and Cl- on the photodegradation kinetics of SAR and GAT. The design layout and data analysis were performed using Design Expert (version 7.1.3, Stat-Ease Inc., Minneapolis, MN) and according to Ferry et al. (17, 19). In brief, the relationship between the photodegradation rate constants (k) and the four variables was evaluated by fitting a full quadratic expression: k ) β0 + β1x1 + β2x2 + β3x3 + β4x4 + β12x1x2 + β13x1x3 + β14x1x4 + β23x2x3 + β24x2x4 + β34x3x4 + β11x12 + β22x22 + β33x32 + β44x42 (1) where x1-x4 represent the levels of the four factors (Table 1), and βx (β0-β44) are the regression parameters to be fitted. As HA and Cl- are prominent water constituents in natural waters, the individual role of HA and Cl- on the photodeg-
radation kinetics of SAR and GAT was examined. To examine the effects of ionic strength and pH on the photoreaction kinetics, Na2SO4 and phosphate buffers were used to adjust ionic strength and pH, respectively. Dark controls were performed under the same conditions. In addition, the solutions of the FQs in pure water and natural waters were preserved in the dark for up to 50 days at room temperature to examine their aqueous stability. Photodegradation experiments and the dark controls were carried out at least in triplicate. Analytical Determinations. An Agilent 1100 HPLC with a DAD was employed to analyze the FQ concentrations, and the abundance of the photodegradation intermediates. The intermediates were enriched by solid phase extraction and then analyzed by an Agilent 1100 LC/MSD and a Thermo Electron LC/MS/MS. Ions were analyzed with a Shimadzu Class-VP ion chromatography. The analytical details are presented in the SI. Biotest for Photoproducts. The bioluminescence inhibition assay using Vibrio fischeri was adopted to indicate the toxicity variation during the FQ photodegradation. The initial concentrations (C0 ) 100 µM) were the same as those in product identification studies. The 15 min bioluminescence assay was carried out following the China standard method GB/T 15441-1995 (41), and a LuminMax-C Luminometer was used to quantify the luminescence intensities. Luminescence inhibition rate was calculated according to Jiao et al. (42).
Results and Discussion Photodegradation Kinetics in Pure Water. In the dark controls, no obvious loss of the eight FQs was observed, indicating the decay by microbiological, thermal, or hydrolytic means was negligible during the photodegradation experiments. Upon preserving the dark control solutions for up to 50 days, the eight FQs decreased by only 1.5-13.7% (SI Table S2), which indicates that these FQs are hydrolytically stable in the studied waters. The FQs absorb the actinic portion of the simulated solar spectrum (Figure 1). When exposed to the simulated solar irradiation, the FQs disappeared quickly. Linear regression of ln(C/C0) vs time (t) showed the photodegradation followed the pseudofirst-order kinetics (r2 > 0.95). The photodegradation rate constants (k), half-lives (t1/2) and r2 values are listed in SI Table S3. In pure water, t1/2 ranged from 15.6 ( 2.6 min for ENR to 140.9 ( 16.5 min for GAT. The determined Φ values for the FQ photolysis in pure water are listed in Table 2. They ranged from (4.72 ( 0.56) × 10-3 for BAL to (6.97 ( 1.41) × 10-2 for ENR. Based on the VOL. 44, NO. 7, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 3. Parameter Estimates and Hypothesis Tests for the Coefficients of the Quadratic Model Fitted to the Photodegradation Data of Sarafloxacin parameter
βx key
β0 β1 β2 β3 β4 β12 β13 β14 β23 β24 β34 β11 β22 β33 β44
intercept Fe(III) NO3– HA Cl– Fe(III)-NO3– Fe(III)-HA Fe(III)-Cl– NO3––HA NO3–-Cl– HA–Cl– Fe(III)2 (NO3–)2 (HA)2 (Cl–)2
FIGURE 2. Effect of pH on the aqueous photodegradation of sarafloxacin. Error bars represent the 95% confidence interval (n ) 3). Φ values and the method of Leifer (43), half-lives were calculated for solar photodegradation of the FQs in pure water and at 45° N latitude. The calculated values ranged from 1.25 min for ENR in midsummer to 58.0 min for BAL in midwinter (Table 2). The t1/2 values are comparable with the determined values under sunlight irradiation (SI Table S4). These are short half-lives compared to other antibiotics, such as sulfa drugs with t1/2 ) 9.2-420 h (39). Thus, the photodegradation can be a central factor in determining the fate of FQs in sunlit surface waters. Photodegradation in Natural Waters and the pH Effect. As shown in SI Table S3, k values for CIP, DAN, and SAR in freshwater and seawater are comparable to those in pure water. However, for the other five FQs, k values in natural waters are evidently lower than those in pure water. The similar or different photodegradable potential in natural waters compared with pure water can be attributed to the integrative effects of the different water characteristics (such as ionic strength and pH) and constituents (e.g., HA and Cl-, the main constituents in freshwater and seawater, respectively) on the photodegradation (22). Thus, the role of ionic strength and pH on the photodegradation was investigated with SAR and GAT as representatives. The addition of Na2SO4 did not influence the k values of SAR and GAT (SI Figure S2). Thus, the photolysis kinetics were independent of Na+, SO42-, or ionic strength, which is consistent with the observations of Prabhakaran et al. (44). With the increase of pH from 5 to 11, k of SAR increased first and then decreased with a peak value at pH 8 (Figure 2). GAT has the similar trend of k variations with pH (SI Table S3). The approximate isoelectric points (pHiso) could be obtained by averaging the dissociation constants (45), pKa1 and pKa2 (SI Table S1). For SAR and GAT, pHiso are approximately equal to 7.4 and 7.6, respectively. At the pHiso, zwitter ions were prevalent for the two antibiotics. Thus, it can be deduced that the zwitter ionic form of SAR and GAT photodegraded faster than the other acidic or basic forms. Different absorption spectra and quantum yields of these forms may account for the pH dependence (28, 33, 46, 47). For the FQs dissolved in freshwater or seawater, the solution pH was closer to their corresponding pHiso values (SI Table S3), so the photodegradation of FQs in the natural waters (pH 7.54-8.37) should be faster than that in pure water (pH 6.38-6.64). However, the photodegradation in the natural waters was slower in most cases (SI Table S3), which implied that water constituents might influence the photodegradation more significantly than pH did. To verify the hypothesis, the effects of the main natural water constituents on SAR and GAT photolysis were studied. Multivariate Effects of HA, Cl-, NO3- and Fe(III) on SAR and GAT Photolysis. The kinetic parameters for the photodegradation of SAR under different conditions in the central composite experiments were listed in SI Table S5. The βx values and the corresponding significance level (p) were 2402
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a
βx × 103 sum of squares × 106 14.00 –1.60 –1.59 –5.18 1.32 0.29 1.80 –1.83 1.41 –0.54 –2.65 –0.56 –0.64 1.34 –0.87
61.2 60.6 644.7 41.6 1.4 52.1 53.3 31.6 4.6 112.2 8.5 11.3 48.9 20.8
p 0.0418a 0.0426a < 0.0001a 0.0864 0.7454 0.0579 0.0553 0.1305 0.5503 0.0087a 0.4194 0.3546 0.0651 0.2139
Tests as significant at the 95% confidence level.
shown in Table 3. For SAR, Fe(III), NO3-, HA, and the interaction between HA and Cl- were significant contributing factors. Equation 1 can be simplified as follows: k ) β0 + β1x1 + β2x2 + β3x3 + β34x3x4
(2)
For eq 2, the coefficient of determination (R2) adjusted by degree of freedom was 0.735. All the βx values in eq 2 are negative, indicating that Fe(III), NO3-, HA and HA-Clinteraction inhibited the photodegradation of SAR. Based on the results of the hypothesis tests (Table 3), HA was the most significant factor inhibiting the photodegradation of SAR, followed by HA-Cl-, Fe(III) and NO3-. According to the sum of squares (Table 3), the inhibiting contribution of HA, HA-Cl-, Fe(III), and NO3- were estimated to be 55.9, 9.7, 5.3, and 5.3%, respectively. For Cl-, p > 0.05, implying that Clhad no significant impact on the photodegradation. The inhibitive effect of HA and the unobvious effect of Cl- were validated by the sole addition of HA or Cl- in the SAR solutions (SI Table S3). As shown in SI Table S6, HA also inhibited the photodegradation of GAT significantly (p < 0.05), followed by NO3(p < 0.1). Thus, whether the similar photodegradation potential of SAR in different waters or the less photodegradable potential of GAT in natural waters than in pure water, was attributed to the integrative effects of pH, HA, and NO3-. The following interpretations can be given for the inhibitive effects of HA, NO3-, Fe(III) and HA-Cl- interaction. HA have strong overlapping absorption spectra with SAR or GAT (SI Figure S3). Thus, HA competitively absorbed actinic photons (λ ) 290-370 nm). The competitive absorption could also be caused by NO3-, and Fe(III) or Fe(III)-aquo complexes near 290 nm (48, 49). Moreover, HA could scavenge the ROS (e.g., •OH and 1O2) and inhibited the possible self-sensitized photolysis of the FQs (17, 22). HA-Cl- interaction might also have the similar mechanisms, competitive photoabsorption and ROS scavenging. Considering the abundance of HA, NO3-, Fe(III), and Cl- in natural waters (SI Table S7), the two mechanisms explained most of the less photodegradable potential of the FQs in fresh water and seawater than in pure water. There were also studies reporting that HA, NO3-, and Fe(III) sensitized the photodegradation of antibiotics, for example, sulfadimethoxine (50), phenicols (20), and lincomycin (51). We cannot preclude the sensitizing effects of these constituents on the FQ photodegradation. However, the experimental results indicated that the inhibiting effects caused by competitive photoabsorption or ROS scavenging were stronger than the sensitization effects.
FIGURE 3. Main transformation pathways for photodegradation of fluoroquinolones in pure water. The photoproducts are labeled “Pm-n”, with m standing for the number of the FQs (Table 2), and n for the number of the intermediates. Photodegradation Types, Products, and Pathways. The addition of isopropanol (•OH quencher) in pure water induced a pronounced retardation of the photodecomposition rate of the eight FQs (SI Figure S4), indicating that their photoreactions involved self-sensitized photo-oxidation via •OH. The addition of NaN3 (•OH and 1O2 quencher) in pure water inhibited the photodegradation of all the FQs except for CIP and DAN (SI Figure S4). For LEV, GAT, DIF, and BAL, the inhibitive effect of NaN3 was more significant than that of isopropanol, suggesting that the four FQs also underwent 1O2-mediated self-sensitized photolysis. The other four FQs might still undergo 1O2-mediated photolysis, for the addition of NaN3 increased the solution pH (SI Table S3) and quickened the FQ photolysis to some extent (Figure 2). For the photodegradation of the eight FQs in pure water, F- and HCOO- were identified as products, and Fconcentrations (CF) were quantified (SI Table S8). Furthermore, 28 main organic intermediates for six FQs were also identified (SI Table S9 and Figure S5). The intermediates of GAT and BAL were not detected, possibly due to their less photostability compared with the parent compounds. Within the 28 intermediates, there were 15 abundant products, for which the responses in ESI (+) MS mode were the most significant. The abundance and
evolvement of the 15 intermediates were shown in SI Figure S6. The photoproducts are labeled “Pm-n”, with m standing for the number of the FQs (Table 2) and n for the number of the intermediates. Except for P2-3, the structures were elucidated for the other 27 products. CIP (P6-1) and SAR (P5-1) were formed from the photodegradation of ENR and DIF, respectively. Intermediates of CIP (P1-2) and SAR (P4-2, P4-3) were produced by the substitution of -F by -OH, which was attributed to the attack of •OH to -F and validated the involvement of •OH in the photoreactions of the FQs. P1-3, P2-4, P5-3, and P6-3 might be generated from oxidation of the piperazine ring by 1O2, proving the participation of 1O2 in the FQ photoreactions. The photodegradation pathways are proposed according to the photoproducts, which involve three main pathways: photoinduced decarboxylation, hydroxylated defluorination, and piperazinyl N4-dealkylation (Figure 3). The three pathways can be prioritized based on CF (SI Table S8) and the product abundance (SI Figure S6). For DAN, LEV, DIF, and ENR, N4-dealkylation was the main pathway, followed by decarboxylation and defluorination. However, for CIP and SAR, decarboxylation and defluorination were comparatively more important due to their outer nonaliphatic N4 in the piperazine moiety. In addition to the three main pathways, VOL. 44, NO. 7, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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some inefficient degradation pathways were also found, such as oxidation and fragmentation of the piperazine ring. These pathways occurred effectively for CIP catalyzed by visiblelight-mediated TiO2 (2), oxidated by manganese oxide (3) and interacting with HOCl (52). Photomodified Toxicities of FQs to Vibrio fischeri. The toxicity evolvement of the 6 FQs during the irradiation was assayed, for which the results are shown in SI Figure S7. The luminescence inhibition rate of the parent FQs to Vibrio fischeri ranged from 16.5 ( 4.8% for SAR to 85.0 ( 4.4% for DIF. The six FQs have a similar toxicity evolvement trend. During the photodegradation, the toxicities first decreased, then increased, and finally decreased, implying the generation of some more toxic intermediates than the parent compound. In natural water bodies, the levels of FQs are usually lower than the concentrations employed in this study (4-10). Thus, the FQ levels may be too low to exhibit their photomodified toxicities. Environmental Significance. This study found exposure to sunlight will facilitate the decay of FQs in the euphotic zone of surface waters. However, the photolytic pathways and the multivariate role of water constituents on FQ photolysis need to be taken into account upon assessing the fate of FQs. Despite the high FQ concentrations employed in the study, the photomodified toxicities urge more concerns on the ecological risk for the class of antibiotics.
Acknowledgments We thank Dr. Guangshui Na for the assistance on LC/MS/MS analysis. This work was supported by the National Basic Research Program (2006CB403302), National Natural Science Foundation (20777010) and Program for Changjiang Scholars and Innovative Research Team in University (IRT0813) of China.
Supporting Information Available Detailed analytical procedures, UV-vis spectra of HA and L-HA, list of photodegradation kinetic data for the FQs, LC/ MS/MS data, biotest results. This material is available free of charge via the Internet at http://pubs.acs.org.
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